EPA/635/R-02/001F

TOXICOLOGICAL REVIEW

OF

BENZENE

(NONCANCER EFFECTS)

(CAS No. 71-43-2) 

In Support of Summary Information on the 

Integrated Risk Information System (IRIS) 

October 2002 

U.S. Environmental Protection Agency 

Washington, DC 
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. 

ii 
CONTENTS—TOXICOLOGICAL REVIEW FOR BENZENE (CAS No. 71-43-2) 

FOREWORD ............................................................... vii

AUTHORS, CONTRIBUTORS, AND REVIEWERS ............................... viii

LIST OF ACRONYMS AND ABBREVIATIONS ................................... ix

EXECUTIVE SUMMARY .................................................... xii

1. INTRODUCTION ..........................................................1

2. CHEMICAL AND PHYSICAL INFORMATION RELEVANT TO ASSESSMENTS .....2

3. TOXICOKINETICS RELEVANT TO ASSESSMENTS ............................3

3.1. ABSORPTION .....................................................3

3.1.1. Gastrointestinal Absorption ...................................3

3.1.2. Respiratory Absorption ........................................4

3.1.3. Dermal Absorption ...........................................5

3.2. DISTRIBUTION ....................................................8

3.2.1. Oral Exposure ...............................................8

3.2.2. Inhalation Exposure ..........................................8

3.2.3. Dermal Exposure ............................................9

3.3. METABOLISM .....................................................9

3.3.1. Metabolic Pathways .........................................10

3.3.2. Requirement for CYP2E1 .....................................10

3.3.3. Toxicity of Benzene Metabolites ...............................12

3.3.3.1 Phenolic Products ...................................12

3.3.3.2. Trans,trans-Muconaldehyde ...........................13

3.3.3.3. Benzene Oxide ......................................14

3.3.4. Species, Route, and Rate Differences ............................14

3.3.5. Induction of CYP2E1 ........................................15

3.3.6. Mechanism of Toxicity .......................................16

3.3.6.1. Formation of Covalent Adducts .........................16

3.3.6.2 Genotoxicity ........................................18

3.3.6.3 Oxidative Stress .....................................18

3.3.6.4. Inhibition of Cytokine Formation .......................19

3.4. ELIMINATION AND EXCRETION ...................................19

3.4.1. Oral Exposure ..............................................19

3.4.2. Inhalation Exposure .........................................20

3.4.3. Dermal Exposure ...........................................21

3.4.4. Other Routes of Exposure .....................................21

3.5. PHYSIOLOGICALLY BASED PHARMACOKINETIC MODELS ...........22

3.6. TOXICOKINETICS SUMMARY ......................................26

4. HAZARD IDENTIFICATION ................................................28

iii
CONTENTS–TOXICOLOGICAL REVIEW FOR BENZENE (CAS No. 71-43-2) 

(continued) 

4.1. STUDIES IN HUMANS .............................................28

4.1.1. Oral Exposure ..............................................28

4.1.2. Inhalation Exposure .........................................28

4.1.2.1. Hematotoxicity ......................................28

4.1.2.2. Neurotoxicity .......................................43

4.1.2.3. Reproductive Toxicity ................................44

4.1.2.4. Developmental Toxicity ..............................47

4.2. ACUTE AND CHRONIC STUDIES IN EXPERIMENTAL ANIMALS .......50

4.2.1. Hematotoxicity .............................................51

4.2.1.1. Oral Exposure—Subchronic Studies .....................51

4.2.1.2. Oral Exposure—Chronic Studies ........................52

4.2.1.3. Inhalation Exposure—Subchronic Studies ................53

4.2.1.4. Inhalation Exposure—Chronic Studies ...................60

4.2.1.5. Effects on Stem Cell Populations .......................63

4.2.1.6. Summary of Principal Hematotoxic Effects ...............77

4.2.2. Reproductive/Developmental Effects ............................79

4.2.2.1. Reproductive Toxicity ................................79

4.2.2.2. Developmental Toxicity ..............................83

4.2.2.3. Summary of Principal Reproductive/Developmental 

Toxicity Effects .....................................95

4.2.2.4. Mechanisms of Developmental and Reproductive Toxicity ...96

4.2.3. Neurotoxicity ..............................................97

4.2.3.1. Oral Exposure ......................................97

4.2.3.2. Inhalation Exposure ..................................98

4.2.3.3. Summary of Neurotoxic Effects .......................101

4.2.4. Immunotoxicity ............................................101

4.2.4.1. Oral Exposure .....................................101

4.2.4.2. Inhalation Exposure .................................103

4.2.4.3. Summary of Immunotoxic Effects ......................104

4.3. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS

AND MODE OF ACTION ..........................................104

4.4. SUSCEPTIBLE POPULATIONS .....................................109

4.4.1. Childhood Susceptibility .....................................109

4.4.2. Gender Differences .........................................110

4.4.3. Genetically Susceptible Populations ............................111

4.5. HAZARD IDENTIFICATION SUMMARY ............................113

5. DOSE-RESPONSE ASSESSMENTS .........................................116

5.1. INHALATION REFERENCE CONCENTRATION (RfC) .................117

5.1.1. Choice of Principal Study and Critical Effect ....................117

5.1.2. Benchmark Dose Modeling ..................................119

5.1.3. RfC Derivation ............................................121

5.1.4. Comparison Analysis Based on the LOAEL .....................123

5.1.5. Comparison Analysis Based on the Ward et al. (1985) Experimental

iv
CONTENTS–TOXICOLOGICAL REVIEW FOR BENZENE (CAS No. 71-43-2) 

(continued) 

Animal Study .............................................123

5.2. ORAL REFERENCE DOSE (RfD) ....................................127

5.2.1. Choice of Principal Study and Critical Effect ....................127

5.2.2. Conversion of Inhalation Exposure to Equivalent Oral Dose Rate ....127

5.2.3. RfD Derivation ............................................129

5.2.4. Comparison Analysis Based on the LOAEL .....................130

5.2.5. Comparison Analysis Based on the NTP (1986) Experimental 

Animal Study .............................................130

5.3 DOSE-RESPONSE SUMMARY .....................................135

6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND 

DOSE-RESPONSE ........................................................141

6.1. HUMAN HAZARD POTENTIAL ....................................141

6.2. DOSE-RESPONSE ................................................144

6.2.1. Inhalation RfC .............................................144

6.2.2. Oral RfD .................................................144

7. REFERENCES ...........................................................146

v 
LIST OF FIGURES

Figure 1. Metabolic pathways for benzene .........................................11

Figure 2. Linear model of ALC data .............................................121

Figure 3. Linear model of HCT data .............................................125

LIST OF TABLES 

Table 1. Hematotoxicity of benzene—occupational exposure ..........................30

Table 2. Median benzene urine metabolites, by exposure category, in a study of 

workers exposed to benzene in Shanghai, China, 1992 ..............................40

Table 3. Comparison of mean peripheral blood counts with standard deviations, by 

exposure status, in a study of workers exposed to benzene in Shanghai, China, 1992 ......41

Table 4. Reproductive toxicity of inhaled benzene in humans ..........................45

Table 5. Developmental toxicity of benzene—humans ...............................48

Table 6. Peripheral blood and hematopoietic effects of benzene in animals—

inhalation exposure .........................................................54

Table 7. Reproductive toxicity of inhaled benzene in test animals ......................80

Table 8. Developmental toxicity of inhaled benzene in test animals .....................84

Table 9. Joint effects of CYP2E1 activity and NQO1 genotype on benzene-induced

hematotoxicity in Chinese Workers ..............................................112

Table 10. Results of BMC modeling of Rothman et al. (1996a) data on benzene and ALC . . 122

Table 11. BMD modeling results of the NTP (1986) male mouse and male rat lymphocyte

counts, with untransformed data ..............................................133

Table 12. BMD modeling results of the NTP (1986) male mouse and male rat lymphocyte 

counts, with transformed dose data ............................................134

Table 13. Summary of uncertainty factors used for deriving the RfC and RfD ............137

Table 14. Summary of RfC and RfD estimates using human and experimental animal data, 

as well as benchmark dose modeling and LOAEL/NOAEL approaches ...............142

vi
FOREWORD 

The National Center for Environmental Assessment-Washington Division, Office of 

Research and Development, has prepared this document on the Toxicological Review of Benzene 

to serve as a source document for updating the noncancer health effects summary on benzene in 

the Integrated Risk Information System (IRIS). 

In the development of this document, the published relevant scientific literature has been 

reviewed, key studies have been evaluated, and summary/conclusions have been prepared so that 

the noncancer health effects from exposure to benzene are qualitatively and quantitatively 

characterized and the derivation of the reference dose (RfD) and reference concentration (RfC) are 

adequately described. The evaluation and review of the noncancer effects of exposure to benzene 

have been conducted under the U.S. Environmental Protection Agency’s standing guidance of 

several relevant risk assessment guidelines dealing with reproductive, developmental, and 

neurotoxic effects. This draft has undergone internal peer review and an expert external peerpanel review in October 1998 as well as a 90-day public comment period. 

The emphasis of this document is a discussion of the noncancer adverse health effects of 

benzene, including the no-observed-adverse-effect levels, the lowest-observed-adverse-effect 

levels, benchmark dose analysis, uncertainty factors, and any other considerations used to develop 

the RfDs and RfCs for benzene. 

vii 
AUTHORS, CONTRIBUTORS, AND REVIEWERS 

The National Center for Environmental Assessment-Washington Division (NCEA-W) of 

the U.S. Environmental Protection Agency’s (EPA’s) Office of Research and Development (ORD) 

was responsible for the preparation of this document. TN & Associates, Inc., under EPA Contract 

No. 68-C6-0024, Work Assignment No. 1-09.1, conducted the literature search and prepared 

preliminary drafts of the document. 

AUTHORS 

David Bayliss, NCEA-W 

Jennifer Jinot, NCEA-W 

Babasaheb Sonawane, NCEA-W 

INTERNAL REVIEWERS 

NCEA/ORD 

Robert Bruce, NCEA-Cin

Margaret Chu, NCEA-W

Eric Clegg, NCEA-W

Anne Jarabek, NCEA-RTP

Carole Kimmel, NCEA-W

Robert McGaughy, NCEA-W

William Pepelko, formerly of NCEA-W

James Rowe, Office of Science Policy

Woody Setzer, National Health and Environmental Effects Research Laboratory-RTP

Other EPA Offices 

Pam Brodowicz, Office of Mobile Sources, Office of Air and Radiation 

Other Federal Offices 

Dr. Nathaniel Rothman, National Cancer Institute, National Institutes of Health, Bethesda, 

Maryland 

Peer Review Panel 

Dr. Donald Gardner, Chairperson, Inhalation Toxicology Associates, Raleigh, NC

Dr. Lynne Haber, Toxicology Excellence for Risk Assessment, Cincinnati, OH

Dr. John Keller, Consultant, Olney, MD

Dr. Michele Medinsky, Consultant, Durham, NC

Dr. Robert Snyder, Department of Pharmacology and Toxicology, Rutgers University, 

Piscataway, NJ

viii 
LIST OF ACRONYMS AND ABBREVIATIONS 

AChE acetyl cholinesterase

ALC absolute lymphocyte count

AML acute myelogenous leukemia

ANOVA analysis of variance

ASA acetylsalicylic acid

AUC area under the curve

BFU-E burst-forming units–erythroid

BMC benchmark concentration

BMCL 95% lower confidence limit on the BMC

BMD benchmark dose

BMDL 95% lower confidence limit on the BMD

BMR benchmark response

CFU-C colony-forming units–culture

CFU-C/tibia colony-forming granulopoietic stem cells

CFU-E colony-forming units–erythroid

CFU-HPP high proliferative potential colony-forming units

CFU-S colony-forming units–spleen

CHO-M-OH 6-hydroxy-trans,trans-2,4-hexadienal

confidence interval 

Con A concanavalin A 

COOH-M-CHO 6-oxo-trans,trans-2,4-hexadienoic acid 

COOH-M-OH 6-hydroxy-trans,trans-2,4-hexadienoic acid 

CYP2E1 cytochrome P450 2E1 

EMG electromyographical

EPA U.S. Environmental Protection Agency 

F344 Fischer 344 

GD gestation day 

ix 

CI 
LIST OF ACRONYMS AND ABBREVIATIONS (continued) 

GM-CFU-C granulocyte/macrophage colony-forming unit–culture 

GM-CFU-G granulocyte/macrophage colony-forming unit 

GPA glycophorin A 

GSH glutathione 

GTP guanosine triphosphate 

HCT hematocrit 

Hgb hemoglobin 

HSDB Hazardous Substance Data Bank 

IC50 Inhibition concentration 50% 

i.p. intraperitoneal 

IL-1 interleukin-1 

IL-1 interleukin-1alpha 

IL-1 interleukin-1beta 

IRIS Integrated Risk Information System 

Kd kilodalton 

LOAEL lowest-observed-adverse-effect level 

LPS lipopolysaccharide 

MA trans,trans-muconic acid 

MCH mean corpuscular hemoglobin 

MCV mean corpuscular volume 

MDS myelodysplastic syndrome 

MPO myeloperoxidase 

MUC muconaldehyde 

NADPH nicotinamide adenine dinucleotide phosphate 

NCEA National Center for Environmental Assessment 

NCEA-W National Center for Environmental Assessment–Washington 

Division 

NK natural killer 

x 
LIST OF ACRONYMS AND ABBREVIATIONS (continued) 

NOAEL no-observed-adverse-effect level

NQO1 NAD(P)H NAD(P)H quinone oxidoreductase

NTP National Toxicology Program

OR odds ratio

PBPK physiologically based pharmacokinetic

PCE polychromatic erythrocytes

PC-PFC polyclonal plaque-forming cells

PCV packed cell volume

PHA phytohemaglutinin

RBC red blood cell

RfC reference concentration

RfD reference dose

SD standard deviation

SE standard error

SRBC sheep red blood cell

TWA time-weighted average

UF uncertainty factor

U.S. EPA U.S. Environmental Protection Agency

WBC white blood cell

YAC yeast artificial chromosome

xi 
EXECUTIVE SUMMARY 

The U.S. Environmental Protection Agency’s (EPA’s, the Agency’s) Office of Mobile 

Sources, Office of Air and Radiation, requested that the National Center for Environmental 

Assessment (NCEA) provide an updated characterization of human health risk from exposure to 

benzene. Recently, EPA published Carcinogenic Effects of Benzene: An Update (EPA/600/P97/001F, April 1998 [U.S. EPA, 1998a]) and Extrapolation of the Benzene Inhalation Unit Risk 

Estimate to the Oral Route of Exposure (NCEA-W-0517, April 1999). These documents serve 

as a source of information regarding the carcinogenic effects of exposure to benzene. 

The scope of the present report is limited to an assessment of the noncancer effects of 

benzene under the Agency’s standing guidance of several relevant risk assessment guidelines 

dealing with reproductive, developmental, neurotoxic, and other noncancer effects, including 

derivation of an oral reference dose (RfD) and inhalation reference concentration (RfC). This 

toxicological review of benzene is to serve as a scientific source document for hazard 

identification and dose-response assessment in updating the noncancer health effects summary 

on benzene in the Integrated Risk Information System (IRIS). 

Benzene, also known as benzol, is widely used as an industrial solvent, as an 

intermediate in chemical syntheses, and as a component of gasoline; hence, the potential for 

human exposure is great. Inhalation exposure is the major route of exposure to benzene, 

although oral and dermal routes are also important. The toxicokinetics (absorption, distribution, 

metabolism, and elimination) of benzene have been studied in humans and experimental animal 

species. Benzene is readily absorbed by both test animals and humans and is distributed among 

several body compartments. The parent compound is preferentially stored in fat, and the relative 

uptake appears to be dependent on the blood perfusion rates of tissues. Metabolism of benzene 

is required for expression of benzene toxicity. 

Evidence indicates that following inhalation exposure to benzene, the major route of 

elimination of unmetabolized benzene in humans is via exhalation. Absorbed benzene is 

metabolized to phenol and muconic acid, followed by urinary excretion of conjugated sulfates 

and glucuronides. Limited data exist on excretion of benzene in humans following dermal 

exposure. Physiologically based pharmacokinetic models have been developed and are being 

improved upon to better define interactions of benzene metabolism, toxicity, and dosimetry. 

These interactions for humans exposed to low concentrations can be assessed only when the 

mode of action is understood at a quantitative level and is incorporated within a physiological 

modeling framework. However, the current models are insufficiently developed and validated to 

allow them to predict with certainty the relationship between metabolism and toxicity of 

benzene. 

Benzene exposure results in adverse noncancer health effects by all routes of administration 

to test animal species. Hematotoxicity has been consistently reported to be the most sensitive 

indicator of noncancer toxicity both in limited studies in humans and experimental animals, with 

bone marrow as the principal target organ. Chronic exposure to benzene results in progressive 

deterioration of hematopoietic function. Whether the hematotoxic and carcinogenic effects of 

benzene are due to a common mechanism has not been established. Although leukocytopenia has 

been consistently shown to be a more sensitive indicator than anemia of benzene toxicity in 

experimental animals, lymphocytopenia has been shown to be even more sensitive than 

xii 
leukocytopenia. A decrease in absolute lymphocyte count (ALC) was observed to be the most 

sensitive indicator of benzene exposure in several epidemiologic studies. 

Benzene has also been shown to produce neurotoxic effects in test animals and humans after 

short-term exposure, but at relatively high concentrations; however, long-term neurotoxicity 

exposure studies are lacking. There is some evidence of adverse reproductive and developmental 

effects due to benzene exposure from human epidemiologic studies, but data are not conclusive 

enough to link low exposure levels to effects. No data from human studies were found to indicate 

that children are more sensitive than adults to benzene toxicity, nor were any significant gender 

differences found. The most frequently observed gender difference in test animals is a greater 

sensitivity of male mice to benzene exposure; in rats, females appear to be more sensitive than 

males. However, in the National Toxicology Program (NTP) study (NTP, 1986) the male was the 

most sensitive sex in the rat. 

Although a large number of epidemiologic and experimental animal studies are available for 

evaluating noncancer health effects, there are few reports of human data with sufficiently reliable 

estimates of exposure to benzene and few long-term, repeated-dose experiments in test animals. 

Many of the epidemiologic studies are also complicated by exposure to other solvents, and some of 

the longer-term animal studies have employed exposure levels that were too high to establish 

reliable no-observed-adverse-effect level (NOAEL) values. 

The epidemiologic occupational inhalation study by Rothman et al. (1996a) was chosen as 

the principal critical study for deriving both an RfC and an RfD. This study showed significant 

reductions in ALC, red blood cells (RBCs), and platelets in a subgroup of Chinese factory workers 

exposed to a median 8-hour time-weighted average (TWA) concentration of 13.6 ppm (43.4 mg/m3

). 

In a subgroup of 11 workers exposed to a median 8-hr TWA concentration of 7.6 ppm (24 mg/m3), 

only ALC was still significantly reduced. Thus this concentration is a lowest-observed-adverseeffect level ( LOAEL) for benzene immunotoxicity in humans. The findings from this study also 

indicate that white blood cells were significantly decreased and the mean corpuscular volume was 

significantly increased in the total exposed group of 44 workers occupationally exposed to a median 

8-hour TWA of 31 ppm (99 mg/m3

) in comparison to an age- and sex-matched control group. These 

dose-response effects are consistent with the well-known hematotoxic effects of benzene in humans 

and experimental animals, as discussed in Sections 4.1.2.1 and 4.2.1 and summarized in Tables 1 

and 6. 

Reduction in ALC was the most sensitive endpoint in the Rothman et al. (1996a) study, and 

the ALC exposure-response data were used in benchmark dose (BMD) modeling to obtain a point of 

departure for the derivation of the RfC. It was necessary to transform the data to obtain an adequate 

fit with the models for continuous data; then, the continuous linear model provided the best fit. In 

the absence of a clear definition for an adverse effect for this continuous endpoint, a default 

benchmark response of one standard deviation change from the control mean response was selected, 

as suggested in EPA’s draft Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b). 

This default definition of a benchmark response for continuous endpoints corresponds to an excess 

risk of approximately 10% for the proportion of individuals below the 2nd percentile (or above the 

98th percentile) of the control distribution for normally distributed effects. The BMD modeling thus 

resulted in a benchmark concentration (BMC) of 13.7 ppm (8-hour TWA) and a BMCL (95% lower 

confidence unit on the BMC) of 7.2 ppm (8-hour TWA). 

xiii 
As suggested in EPA’s draft Benchmark Dose Technical Guidance Document, the BMCL 

was chosen as the point of departure for the RfC derivation. After converting to mg/m3 and 

adjusting for continuous exposure, a BMCLADJ of 8.2 mg/m3 was obtained. Dividing this value by 

an overall uncertainty factor (UF) of 300 yields a chronic inhalation RfC of 3 × 10-2 mg/m3

. Because 

the BMC is considered to be an adverse-effect level, an effect-level extrapolation factor analogous to 

the LOAEL-to-NOAEL UF was used. EPA is planning to develop guidance for applying an effectlevel extrapolation factor to a BMD. In the interim, a factor of 3 was used in this analysis. 

Additional factors of 10 for intraspecies variability, 3 for subchronic-to-chronic extrapolation, and 3 

for database deficiencies—due to the absence of a two-generation reproductive/developmental 

toxicity study for benzene—comprise the remainder of the 300 composite UF. 

For comparison, an RfC was also calculated from the LOAEL of 7.6 ppm (8-hour TWA) 

from Rothman et al. (1996a). After converting to mg/m3

 and adjusting for continuous exposure, a 

LOAELADJ of 8.7 mg/m3

 was obtained. Dividing this value by an overall UF of 1000 yields an RfC 

of 9 × 10-3 mg/m3. The UF of 1000 was based on factors of 10 to account for the use of a LOAEL 

because of the lack of an appropriate NOAEL, 10 for intraspecies variability, 3 for subchronic-tochronic extrapolation, and 3 for database deficiencies. This result of 9 × 10-3 mg/m3

 based on the 

LOAEL is in good agreement with the result of 3 × 10-2 mg/m3 based on the BMCL. 

Support for this chronic inhalation RfC has been provided by the experimental animal study 

of Ward et al. (1985). The subchronic inhalation study of Ward et al. (1985) was selected as a 

supporting study because it was the experimental animal study with the longest inhalation exposure 

duration and it provided good dose-response data. The exposure-response relationships for the 

different hematologic endpoints in male mice (the most sensitive sex/species in this study) were 

modeled using a BMD modeling approach, and decreased hematocrit (i.e., volume percentage of 

erythrocytes in whole blood) was chosen as the critical effect. The continuous linear model was 

selected because it provided the best fit. As above, a default level of one standard deviation change 

from the control mean response was used as the benchmark response. A BMC of 100.7 ppm and a 

BMCL of 85.0 ppm were obtained. 

The BMCL was chosen as the point of departure for the RfC derivation. After converting to 

mg/m3

 and adjusting for continuous exposure, a BMCLADJ of 48.5 mg/m3 was obtained. As 

discussed above, a UF of 3 is used as an effect-level extrapolation factor, analogous to a LOAEL-toNOAEL UF, because the BMC is considered an adverse-effect level. In addition, the standard UFs 

of 3 for interspecies extrapolation for inhalation studies and 10 for intraspecies variability are 

applied. A UF of 3 for database deficiencies is used, as above. Finally, a partial UF of 3 was used 

to extrapolate from subchronic to chronic exposure. Dividing the BMCLADJ by the overall 

uncertainty factor of 1000 yields a chronic inhalation RfC of 5 × 10-2 mg/m3

. This value is in good 

agreement with the RfC of 3 × 10-2 mg/m3

 based on BMD modeling of the ALC data from the 

Rothman et al. (1996a) human study. 

Similarly, for comparison purposes, a chronic inhalation RfC can be derived from the 

NOAEL of 30 ppm observed for hematologic effects in the Ward et al. (1985) study. First, the 

NOAEL is converted to mg/m3

 and adjusted to equivalent continuous exposure, yielding 17.1 

mg/m3

. UFs are identical to those employed above, except that no NOAEL-to-LOAEL UF is used; 

thus, the overall UF is 300. Dividing 17.1 by 300 results in an RfC of 6 × 10-2 mg/m3

. This value is 

also in good agreement with the RfC derived from the Rothman et al. (1996a) study. 

xiv 
To derive the RfD, the same BMCL of 7.2 ppm (8-hour TWA) for the default benchmark 

response of one standard deviation change from the control mean response for the critical effect of 

reduced ALC in Rothman et al. (1996a) was used as the point of departure. After converting the 

units, correcting for continuous exposure, and adjusting for the route-to-route extrapolation from 

inhalation to oral exposure, a BMCLADJ-equivalent oral dose rate of 1.2 mg/kg/day was obtained. 

Dividing the BMCLADJ-oral of 1.2 mg/kg/day by an overall UF of 300 yields a chronic oral RfD of 4 

× 10-3 mg/kg/day. As above, a UF of 3 is used as an effect-level extrapolation factor, analogous to 

the LOAEL-to-NOAEL UF, because the BMC is considered to be an adverse-effect level. 

Additional factors of 10 for intraspecies variability, 3 for subchronic-to-chronic extrapolation, and 3 

for database deficiencies comprise the remainder of the 300 composite UF. 

For comparison, an RfD was also calculated from the LOAEL of 7.6 ppm (8-hour TWA) 

from Rothman et al. (1996a). After unit conversion, correction for continuous exposure, route-toroute extrapolation, and division by a combined UF of 1000, an RfD of 1 × 10-3 mg/kg/day was 

derived from the LOAEL. The combined UF of 1000 was based on factors of 10 to account for 

using a LOAEL because of the lack of an appropriate NOAEL, 10 for intraspecies variability, 3 for 

subchronic-to-chronic extrapolation, and 3 for database deficiencies. This RfD value of 1 × 10-3 

mg/kg/day is in good agreement with the value of 4 × 10-3 mg/kg/day calculated from the BMCL. 

For further comparisons, RfDs were calculated on the basis of the NTP’s chronic 

experimental animal study (NTP, 1986). In this study, Fischer 344 (F344) rats and B6C3F1 mice of 

both sexes were administered benzene by gavage, and blood was drawn from subgroups of animals 

at various time points. This study identified a LOAEL of 25 mg/kg for leukopenia and 

lymphocytopenia in female F344 rats and male and female B6C3F1 mice; no NOAEL was 

identified. Reduction in lymphocyte count was selected as the critical effect, and the dose-response 

relationships at different time points were modeled using a BMD modeling approach. As above, a 

default benchmark response of one standard deviation change from the control mean was used to 

define the BMDs. The modeling results suggested that the male rat was the most sensitive 

sex/species in this NTP gavage bioassay, and a BMDL (95% lower confidence limit on the BMD) of 

1 mg/kg was selected as the point of departure for deriving the RfD. 

The BMDL was first adjusted for exposure 7 days/week and then divided by a composite UF 

of 1000. This composite UF reflects a UF of 3 used as an effect-level extrapolation factor, 

analogous to LOAEL-to-NOAEL UF, because the BMD is considered to be an adverse effect; a UF 

of 10 for interspecies extrapolation for oral studies; a UF of 10 for intraspecies variability; and a UF 

of 3 for database deficiencies. The resulting RfD was 7 × 10-4 mg/kg/day, which is in reasonably 

good agreement (within an order of magnitude) with the RfD of 4 × 10-3 mg/kg/day derived from the 

Rothman et al. (1996a) human inhalation study. 

An RfD was also derived from the LOAEL of 25 mg/kg observed in the NTP (1986) study. 

The LOAEL was adjusted to a continuous exposure level of 17.9 mg/kg/day and then divided by a 

UF of 3000 to derive an RfD of 6 × 10-3 mg/kg/day. The combined UF of 3000 is based on factors 

of 10 for the absence of a NOAEL, 10 for interspecies extrapolation, 10 for intraspecies variability, 

and 3 for database deficiencies. This value of 6 × 10-3 mg/kg/day is in good agreement with the 

value 4 × 10-3 mg/kg/day derived from the Rothman et al. (1996a) human study. 

xv 
1. INTRODUCTION 

The overall purpose of this document is to provide a toxicological review of noncancer 

health effects of exposure to benzene, including the no-observed-adverse-effect levels 

(NOAELs), the lowest-observed-adverse-effect levels (LOAELs), benchmark dose (BMD) 

modeling, and uncertainties in derivation of an inhalation reference concentration (RfC) and an 

oral reference dose (RfD). This toxicological review document may serve as the scientific basis 

for establishing a national air quality standard for ambient air exposure to benzene by the U.S. 

Environmental Protection Agency’s (EPA’s) Office of Mobile Sources, Office of Air and 

Radiation, under the Clean Air Act Amendments of 1990, Subchapter II (42 U.S.C. 7521-7590) 

for hazardous air pollutants. 

This document presents background and justification for the noncancer health hazard 

and dose-response assessment summaries for benzene exposure in EPA’s Integrated Risk 

Information System (IRIS). IRIS summaries include an oral RfD, an inhalation RfC, and a 

carcinogenicity assessment. 

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 milligrams per kilograms per day (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 it provides a continuous inhalation exposure estimate. The inhalation RfC considers 

toxic effects for the respiratory system (portal-of-entry effects) and for systems peripheral to the 

respiratory system (extrarespiratory or systemic effects). It is generally expressed in units of 

milligrams per cubic meter (mg/m3

). The RfC and RfD do not represent a sharp dividing line 

between safe and unsafe. In establishing the RfC and RfD, all relevant biologically significant 

noncancer health effects in the published literature were reviewed and considered. 

Development of these hazard identification and dose-response assessments for benzene 

has followed the general framework for risk assessment, as set forth by the National Research 

Council (1983). EPA guidelines that were used in the development of this assessment include 

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 Neurotoxicity Risk Assessment (U.S. EPA, 

1998b), Proposed Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1996a), Draft Revised 

Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999b), Guidelines for Reproductive 

Toxicity Risk Assessment (U.S. EPA, 1996b), Risk Assessment of Guidelines 1986 (U.S. EPA, 

1987), Recommendations for and Documentation of Biological Values for Use in Risk 

Assessment (U.S. EPA, 1988), Methods for Derivation of Inhalation Reference Concentrations 

and Application of Inhalation Dosimetry (U.S. EPA, 1994), Use of the Benchmark Dose 

Approach in Health Risk Assessment (U.S. EPA, 1995), and Benchmark Dose Technical 

Guidance Document (external review draft; U.S. EPA, 2000b). 

Literature search strategies employed for this compound were based on the Chemical 

Abstracts Registry Number (CASRN) and at least one common name. At a minimum, the 

1 
following databases were searched: RTECS, HSDB, TSCATS, CCRIS, GENETOX, EMIC, 

EMICBACK, DART, ETICBACK, TOXLINE, CANCERLINE, MEDLINE, and MEDLINE 

backfiles. Any pertinent scientific information submitted by the public to the IRIS Submission 

Desk was also considered in the development of this document. 

2. CHEMICAL AND PHYSICAL INFORMATION RELEVANT TO ASSESSMENTS 

Benzene also is known as benzol. Some relevant physical and chemical properties of 

benzene are listed below. 

• CASRN: 71-43-2 

• U.S. EPA Hazardous Waste No.: U019 (in commercial product) 

F005 (nonspecified source, spent solvent) 

• Hazardous Substance Data Bank (HSDB, 1997) No.: 2554 

• Synonyms: annulene, benzeen (Dutch), benzen (Polish), benzine, benzol, benzole, 

benzolo (Italian), bicarburet of hydrogen, coal naphtha, cyclohjexatriene, fenzen (Czech), 

mineral naphtha, motor benzol, NCI-C55276, phene, phenyl hydride, pyrobenzol, 

pyrobenzole 

• Registered trade name: Polystream 

• Empirical formula: C6H6 

• Molecular weight: 78.11 (Budvari, 1989) 

• Vapor pressure: 75 mm Hg at 20°C (NFPA, 1994) 

• Water solubility: 1750 mg/L at 25°C (Banerjee et al., 1980). Miscible with ethanol, 

ethyl ether, acetone, and chloroform. 

• Partition coefficients: Log Kow: 2.13 (Hansch and Leo, 1985) 

Log Koc: 1.8 to 1.9 (HSDB, 1997) 

• Henry’s law constant: 5.5 × 10-3 atm-m3/mol (Mackay and Leinonen, 1975) at 25°C 

• Conversion factor: 1 ppm = 3.24 mg/m3 at 20°C; 1 mg/m3 = 0.31 ppm; 1 mg/L = 313 ppm 

• Melting point: 5.5oC (HSDB, 1997) 

• Boiling point: 80.1oC at 760 mm Hg (HSDB, 1997) 

• Odor: Aromatic (HSDB, 1997) 

Lower 0.84 ppm (HSDB, 1997) 

Upper 53 ppm (HSDB, 1997) 

2 
• Taste threshold: 0.5 to 4.5 mg/L (HSDB, 1997) 

• PEL (permissible exposure limit) TWA = 1 ppm (OSHA, 1987) 

• TLV (threshold limit value) TWA = 10 ppm (ACGIH, 1996) 

3. TOXICOKINETICS RELEVANT TO ASSESSMENTS 

The absorption, distribution, metabolism and excretion of benzene have been intensively 

investigated in several experimental animal species and in humans. Benzene is readily absorbed 

from oral and inhalation exposures. Dermal absorption is also rapid; however, quantitatively, 

dermal absorption is very low due to rapid evaporation from skin. Benzene is rapidly distributed 

throughout the body after exposure by all routes, and accumulation in fatty tissues is observed. 

Metabolism of benzene is necessary for the expression of the characteristic hematotoxic and 

carcinogenic effects of benzene. Despite extensive research, no single metabolite has been 

identified as responsible for all the toxic effects of benzene, and the weight of evidence points 

toward an interaction of several metabolites. At low doses, benzene is rapidly metabolized and 

excretion occurs primarily as conjugated metabolites; however, at higher doses metabolic 

pathways become saturated, and exhalation of unmetabolized benzene is observed to be the 

primary route of excretion. Therefore, extrapolation from results observed at high doses 

underestimates the potential toxic effects of low doses. The pathways of benzene metabolism 

appear to be qualitatively similar across species. Quantitative differences in metabolism among 

animal species are observed, however, and no good model for human metabolism has been 

established. Thus, despite the development of several physiologically based pharmacokinetic 

(PBPK) models, extrapolation of results from laboratory animal results to humans has proved 

difficult. 

3.1. ABSORPTION 

Benzene is readily absorbed by both test animals and humans from inhalation, oral, and 

dermal exposures (U.S. EPA, 2000a). 

3.1.1. Gastrointestinal Absorption 

Although only limited data are available on absorption of benzene in humans by the oral 

route, accidental or intentional poisoning case studies indicate that benzene is readily absorbed 

(Thienes and Haley, 1972). 

Nearly complete absorption of orally administered benzene has been demonstrated in 

laboratory animal studies. Parke and Williams (1953) performed a mass-balance study using 

radiolabeled 14C-benzene with gavage exposures in rabbits at doses of 340–500 mg/kg. 

Calculation of the mass balance 2–3 days after exposure showed that 84–89% of the 

administered 14C could be accounted for as metabolites, CO2, and exhaled unchanged benzene. 

Thus, approximately 90% of the dose was absorbed. Sabourin et al. (1987) similarly found that 

gastrointestinal absorption of 14C-benzene administered to rats and mice at a dose range of 0.5 to 

150 mg/kg was greater than 97%. Although this study used corn oil as a vehicle, it is reasonable 

3 
to assume that oral absorption from water solutions would also be close to 100%. One could 

expect that the presence of food in the stomach may delay absorption and an empty stomach may 

enhance it. 

The presence of soil had only minor effects on absorption of orally administered 14Cbenzene (Turkall et al., 1988). Male rats were gavaged with 5% aqueous solutions of gum 

acacia with either benzene alone or with benzene preadsorbed onto a sandy soil or a clay soil. 

Presence of soil had no significant effect on the absorption of benzene, as less than 2% of the 

administered dose was excreted in the feces within 48 hours. Adsorption of benzene to soil 

apparently increased the rate of gastrointestinal absorption, as the peak plasma benzene 

concentrations were higher in both soil treatments, and the time to peak was reduced in the sandy 

soil. The presence of both soils also increased the proportion of 14C excreted in the urine, as 

opposed to excretion as unmetabolized benzene in expired air. Increased excretion in the urine 

also indicated that the proportion of the benzene subject to metabolism was increased. Phenol 

was the primary metabolite detected in the acid-hydrolyzed urine in all treatment groups, 

followed by hydroquinone, catechol, and benzenetriol. No significant differences in the 

distribution of metabolites among the treatments were found. Because the urine extracts were 

acid-hydrolysed, conjugation products were not determined. Gastrointestinal absorption was 

apparently rapid and efficient, even in the presence of soil. 

3.1.2. Respiratory Absorption 

There is a significant experimental database on the respiratory absorption of benzene in 

humans (Nomiyama and Nomiyama, 1974; Pekari et al., 1992; Srbova et al., 1950; Yu and 

Weisel, 1998). Srbova et al. (1950) examined absorption of benzene by 23 human subjects 

exposed to a range of concentrations, from 47 to 100 ppm (150 to 320 mg/m3

), for 2–3 hours. 

Absorption was greatest in the first 5 minutes of exposure (70–80%) but declined rapidly over 

the next 15 minutes and varied between 20 and 60% after 1 hour and between 20 and 50% after 2 

hours of exposure. Considerable variability between individuals was noted. Nomiyama and 

Nomiyama (1974) determined both retention and uptake of benzene by three female and three 

male subjects, 18–25 years of age, exposed to 52–62 ppm (166–198 mg/m3

) for 4 hours. 

Exhaled air was sampled every hour to determine respiratory excretion of benzene. Retention 

declined from approximately 50% in the first hour and stabilized at 30% after 3 hours. 

Respiratory uptake averaged 47%, with excretion of 17%. No significant differences between 

males and females were reported; however, as noted, only three subjects of each sex were 

examined. 

Pekari et al. (1992) studied respiratory absorption of benzene in three males exposed to 

1.7 and 10 ppm (5 and 39 mg/m3

) benzene for 4 hours each. Absorption of benzene was 

determined by measuring differences in concentration between inhaled and exhaled air. The 

average absorption was 52% at the low concentration and 48% at the high concentration. In a 

recent study, Yu and Weisel (1998) measured the uptake of benzene from sidestream tobacco 

smoke by three female subjects. Benzene concentrations were measured in inhaled and exhaled 

air. Smoke was generated by burning cigarettes, resulting in a variable benzene concentration 

ranging from 32 to 69 ppm (102 to 220 mg/m3). Absorption in eight experiments averaged 64%, 

with a range of 48–73%. The exposure periods were either 30 or 120 minutes, but no significant 

decrease in absorption with the longer exposure period was observed. 

4 
Considered together, these studies indicate that the respiratory absorption of benzene in 

humans is approximately 50%. The observed decline in absorption with increasing exposure 

time is apparently due to respiratory excretion of unmetabolized benzene. Animal studies 

indicate that the metabolism of benzene is saturated at exposure concentrations in excess of 10 

ppm (32 mg/m3

). Thus, with the exception of the study by Pekari et al. (1992), benzene 

metabolism may have been saturated in the human respiratory absorption studies. This would 

have been expected to lead to greater respiratory excretion of unmetabolized benzene. Because 

absorption at low dose levels is most relevant for derivation of an RfD, the estimates of 48–52% 

retention derived by Pekari et al. (1992) are the most relevant values for evaluating absorption in 

humans. 

A number of test animal studies have also been conducted on the absorption of benzene 

from inhalation exposure. Schrenk et al. (1941) noted a linear relationship between benzene 

concentration (200–1300 ppm [639–4153 mg/m3

]) and the equilibrium concentration of benzene 

in the blood of dogs. A steady-state blood level was attained within 30 minutes at these 

exposure concentrations. Exposure concentration appears to affect the retention of inhaled 

radioactivity, as demonstrated by Sabourin et al. (1987) in rats and mice. The retention of 

benzene by rats and mice during a 6-hour exposure decreased as exposure concentration 

increased: 33 ± 6% to 15 ± 9% for rats, and 50 ± 1% to 10 ± 2% for mice, as exposure 

concentration increased from 26 to 2600 mg/m3

 (10 to 1000 ppm). This study also showed 

species variability in the uptake and retention of inhaled benzene. At all exposure 

concentrations, uptake was higher in mice. At exposure concentrations below 350 mg/m3

, mice 

retained approximately 50% more radioactivity per kilogram body weight than did rats, but there 

was no significant difference at the highest (2500 mg/m3) concentration. In general, mice 

inhaled greater amounts of benzene per kilogram body weight due to their higher relative minute 

volume per kilogram than other species. 

Henderson (1996) reviewed species differences in absorption and retention of benzene 

given by either the oral or the inhalation route. Mice have both a higher respiratory rate and a 

faster rate of metabolism for benzene than either rats or monkeys. After a 6-hour exposure to 

7–10 ppm benzene, mice retained 20% of the dose, compared with 3–4% for rats and monkeys. 

At higher exposure levels the metabolism becomes saturated and a sevenfold increase in 

exposure concentration, between 130 and 925 ppm, results in only a threefold increase in 

metabolism. When oral doses exceed the capacity for benzene metabolism, greater 

concentrations are then exhaled. 

3.1.3. Dermal Absorption 

Studies of both humans and experimental animals indicate that benzene is rapidly 

absorbed through the skin from both liquid and vapor phases. The percentage of absorption of 

the applied doses is generally higher in experimental animals than in humans. The percutaneous 

absorption of benzene has been studied in humans (Franz, 1984) and in laboratory animals 

(Maibach and Anjo, 1981; Franz, 1983; Susten et al., 1985). Dermal absorption is minimal when 

compared with inhalation or oral absorption; this is due in large part to benzene volatilizing 

rapidly from the skin. If the absorption is based on the amount applied to the skin without 

accounting for volatilization losses, then percentage absorption figures are low and usually less 

5 
than 1%. Although this represents realistic exposure conditions for most situations, absorption 

can be underestimated in situations where contact with the benzene source is maintained for a 

prolonged period. A substantial portion of the absorbed benzene is excreted through the lungs. 

Thus, experiments that measure excretion only in urine and feces substantially underestimate 

true absorption. 

Franz (1984) studied both in vivo and in vitro dermal absorption of benzene in humans. 

A dose of 0.0026 mg/cm2

 of 14C-benzene was applied to the ventral forearm skin of four 

volunteers. Volatilization was rapid, as no free benzene on the skin was observed after 30 

seconds. Absorption of the applied dose was estimated by urinary excretion of 14C for 36 hours. 

No correction was made for volatilization, but a correction was made for the fact that not all 

absorbed benzene is excreted in the urine. Correction was determined by the fraction of 14C from 

a subcutaneous injection in rhesus monkeys that was not excreted in the urine. Excretion was 

rapid, with 80% of the urinary excretion occurring by 8 hours after exposure. Total absorption 

was estimated to be 0.05% of the applied dose. Absorption using in vitro techniques with human 

skin was estimated to be 0.1% at a similar dose rate. Absorption in vitro increased linearly with 

dose level and also as a function of exposure time. 

Laitinen et al. (1994) studied occupational exposure to benzene in eight car mechanics in 

Finland. Blood samples were taken 3–9 hours after exposure. The approximated benzene 

concentration in blood corresponding to the time point of 16 hours after exposure showed much 

higher levels of exposure than expected, based on corresponding air concentrations in the 

workplace. Actual breathing zone concentrations of benzene varied from the Finish detection 

limit of 0.2 cm3

/m3 to 3.7 cm3/m3, depending on whether it was in unleaded or leaded gasoline. 

Comparison of measured blood concentrations to predictions based on air measurements 

suggested that dermal exposure could have accounted for 68% of exposure. These mechanics 

had direct dermal contact with gasoline during the frequent changing of filters and fuel pumps. 

This suggests far more dermal exposure to benzene than exposure via the inhalation route. 

In a series of experiments conducted in a residence with a benzene-contaminated water 

supply, Lindstrom et al. (1993) estimated, by modeling, that a dose of 281 g benzene would be 

absorbed during a 20-minute shower, with 60% derived from dermal absorption and 40% from 

inhalation absorption, based on results from one individual only. The validity of this observation 

is highly questionable. 

Several dermal absorption studies have also been conducted with experimental animals. 

In rhesus monkeys, minipigs, and hairless mice, dermal absorption was < 1% following a single 

direct application of liquid benzene (Franz, 1984; Maibach and Anjo, 1981; Susten et al., 1985). 

Absorption was rapid, with the highest urinary excretion observed in the first 8 hours following 

exposure (Franz, 1984; Susten et al., 1985). Multiple applications, as well as application to 

cellophane tape-stripped skin resulted in greater skin penetration (Maibach and Anjo, 1981). It 

may be noted the percent absorption of the applied dose of benzene in each of these test animals 

was approximately twofold to threefold higher than that of humans. 

Dermal uptake of benzene from aqueous solutions is an important parameter for 

evaluating the risk due to exposure to benzene-contaminated ground or surface waters, for 

example, during showering, bathing, or swimming. Morgan et al. (1991) compared dermal 

6 
uptake of undiluted benzene with one-third, two-thirds, and completely saturated aqueous 

solutions of benzene. A capped skin depot (3.1 cm2 surface area) was glued to the shaved skin of 

male Fischer 344 (F344) rats, and 2 mL of benzene or aqueous solutions were used. Blood 

samples were collected via an implanted jugular catheter at 0, 0.5, 1, 2, 4, 8, 12, and 24 hours 

after initiation of exposure. The blood concentration during exposure to undiluted benzene 

continued to rise throughout the 24-hour exposure and reached a level of 24.2 mg/L. Dermal 

absorption from aqueous solutions was rapid, and peak blood levels of benzene were reached 

after 1 hour. Benzene could not be detected in blood from the one-third saturated solution; it 

reached a peak level of approximately 0.18 mg/L in the two-thirds saturated treatment and 0.33 

mg/L in the saturated treatment. Benzene was essentially completely absorbed, with less than 

1% remaining in solution after 24 hours. Benzene was preferentially absorbed, because the 

volume of aqueous benzene solution absorbed did not differ significantly from absorption of 

distilled water. Benzene was rapidly and completely absorbed from aqueous solutions. 

Tsuruta (1989) reported that dermal absorption of benzene increased linearly with dose in 

hairless mice exposed to benzene vapors (the mice were attached to respirators to avoid 

inhalation exposure). The dermal absorption rates at exposure concentrations of 200, 1000, and 

3000 ppm (639, 3195, and 9584 mg/m3

) were 4.11, 24.2, and 75.5 mmol/cm2/hour, respectively. 

This is equivalent to an absorption rate of 0.31, 1.89, and 5.90 μg/cm2

/hour, respectively. The 

skin absorption coefficient was 0.619 cm/hour. Using the mouse dermal absorption data and 

human occupational exposure data, Tsuruta estimated that skin absorption of benzene by humans 

would be 3.7% that of inhalation exposure at the same concentration. 

Permeability constants for dermal absorption of benzene vapors were also estimated by 

McDougal et al. (1990). In this study, rats were supplied air through latex masks and exposed to 

benzene vapor at a concentration of 40,000 ppm (127,787 mg/m3

) for 4 hours. Blood 

concentration was monitored at 0.5, 1, 2, and 4 hours. A PBPK model was used to estimate the 

permeability of the vapor in rat skin as well as human skin; the rat and human permeability 

constants were estimated as 0.15 and 0.08 cm/hour, respectively. On the basis of these findings, 

dermal exposure studies in rats probably provide a conservative estimate of the dermal 

absorption of benzene by humans. In an in vitro experiment using F344 rat skin, Mattie et al. 

(1994) determined a skin:air partition coefficient of 35 for benzene at 203 ppm (649 mg/m3

), 

with an equilibration time of 4 hours. 

Adsorption of benzene to soils resulted in a small decrease in dermal absorption when 

compared with free benzene (Skowronski et al., 1988). Radiolabeled 14C-benzene was applied to 

the shaved skin of male Sprague-Dawley rats either as free benzene or mixed with a sandy soil 

(4.4% organic matter) or a clay soil (1.6% organic matter). The application area was occluded 

by gluing a shallow glass cover onto the skin prior to application. Statistically significant 

decreases in the area under the plasma-concentration time curve were observed for both soils, 

but the effect of the clay soil was greater. No statistically significant differences in the 

absorption or elimination half-lives among treatments were observed. The greatest excretion of 

14C was observed in the urine (45–86% of applied dose), with lesser amounts eliminated by 

respiratory excretion (6–13%). Excretion in the feces was 0.2% or less. These values were 

corrected for volatilization losses during administration of 67%, 59%, and 39% of applied 

radioactivity for the free benzene, sandy soil, and clay soil treatments, respectively. Although 

these experiments suggested that soil may reduce benzene absorption, the use of occlusion of the 

7 
exposed skin is not realistic for the dermal exposure scenarios used in risk assessment. Thus, the 

small decreases in uptake have little meaning for evaluation of risks from soil-absorbed benzene. 

3.2. DISTRIBUTION 

There were no studies reported regarding the distribution of benzene in humans after 

either oral or dermal exposure. However, the animal data available provided some useful 

information. Both human and animal data were available following inhalation exposure. 

3.2.1. Oral Exposure 

Low et al. (1989, 1995) followed the distribution of 14C-benzene administered by oral 

gavage in female Sprague-Dawley rats. This study sought to determine whether the distribution 

of benzene or benzene metabolites could explain the observations of tissue-specific induction of 

solid tumors in the Zymbal gland, oral and nasal cavities, and mammary glands associated with 

chronic oral exposure (NTP, 1986). Following oral doses of 0.15 or 1.5 mg/kg in olive oil, the 

concentrations of 14C in various organs could be assigned to three groups, with the highest 

concentrations in liver and kidneys, intermediate concentrations in the blood, and lowest 

concentrations in the Zymbal gland, nasal cavity tissue, oral cavity tissue, mammary glands, and 

bone marrow. At dose levels of 15 mg/kg or higher, disproportionate increases were found in 

mammary glands and bone marrow. At the 0.15 mg/kg dose level, all of the 14C activity found in 

the tissues or blood after 1 hour appeared as benzene metabolites, indicating that first-pass 

metabolism in the liver was very efficient. A high proportion of the metabolites extracted by 

ethyl acetate from the Zymbal gland and the nasal cavity tissue 1 hour after administration were 

present as unidentified metabolites. Similarly, a high proportion of the water-soluble metabolites 

from these tissues were also unidentified peaks that did not correspond to known metabolites of 

benzene. Elimination of benzene-derived 14C in all organs was biphasic. Benzene did not 

accumulate in the Zymbal gland; within 24 hours after administration, radiolabel derived from 

14C-benzene in the Zymbal gland constituted less than 0.0001% of the administered dose. Thus, 

preferential accumulation of benzene or metabolites in the Zymbal gland does not account for 

tissue-specific tumor induction in this organ. 

3.2.2. Inhalation Exposure 

Results from test animal studies indicate that absorbed benzene is distributed throughout 

several body compartments. The parent compound is preferentially stored in the fat, although 

the relative uptake in tissues also appears to be dependent on the perfusion rate of tissues by 

blood. Steady-state benzene concentrations in rats exposed via inhalation to 1600 mg/m3

 (500 

ppm) for 6 hours were blood, 1.2 mg%; bone marrow, 3.8 mg%; and fat, 16.4 mg% (Rickert et 

al., 1979). Benzene was also found in the kidney, lung, liver, brain, and spleen. Levels of the 

benzene metabolites phenol, catechol, and hydroquinone were higher in the bone marrow than in 

blood, with phenol being eliminated more rapidly than catechol or hydroquinone after exposure. 

Ghantous and Danielsson (1986) exposed pregnant mice to a benzene concentration of 6400 

mg/m3

 (2000 ppm) for 10 minutes and found benzene and its metabolites in lipid-rich tissues 

such as brain and fat as well as in perfused tissues such as liver and kidney. Benzene was also 

found in the placenta and fetuses immediately following exposure. 

8 
Benzene was rapidly distributed throughout the bodies of dogs exposed via inhalation to 

concentrations of 800 ppm (2556 mg/m3) for up to 8 hours per day for 8–22 days (Schrenk et al., 

1941). Fat, bone marrow, and urine contained about 20-fold the concentration of benzene in 

blood; benzene levels in muscles and organs were onefold to threefold that in blood, and 

erythrocytes contained about twice the amount of benzene found in plasma. 

Studies in pregnant mice demonstrated that after inhalation exposure, 14C-benzene 

crossed the placenta. Volatile radioactivity (unmetabolized benzene) was observed in the 

placenta and fetuses immediately after and up to 1 hour after exposure (Ghantous and 

Danielsson, 1986). Nonvolatile metabolites were also detected in the fetus, but at lower levels 

than in maternal tissues. The label peaked in fetal tissues 30 minutes to 1 hour after inhalation, 

similar to the peak in maternal tissues. No firmly tissue-bound metabolites of benzene were 

detected in the fetal tissues in late gestation, indicating that the mouse fetus did not have the 

ability to form the reactive metabolites. 

Sato et al. (1975) compared elimination kinetics of benzene in men and women of similar 

ages. Exposure was for 2 hours at 25 ppm. The level of benzene in the blood and the end-tidal 

air was different for males and females. The shape of the decay curve was significantly steeper 

in the males. The authors attributed these results to the higher fat content of females. 

3.2.3. Dermal Exposure 

No relevant studies were found regarding distribution in humans following dermal 

exposure to benzene, and only one study in animals was identified, providing minimal 

information on the distribution of benzene via the dermal route. 

A study of male rats treated dermally with 0.004 mg/cm2

 of 14C-benzene, with and 

without 1 g of clay or sandy soil, revealed soil-related differences in tissue distribution following 

treatment. The 14C activity (expressed as a percentage of initial dose per gram of tissue) 48 

hours after treatment with soil-absorbed benzene was greatest in the treated skin 

(0.059–0.119%), followed by the kidney (0.024%) and liver (0.013–0.015%), in both soil 

groups. In the pure-benzene group, the kidney contained the largest amount of radioactivity 

(0.026%), followed by the liver (0.013%) and treated skin (0.11%) (Skowronski et al., 1988). In 

all three groups, less than 0.01% of the radioactivity was found in the following tissues: 

duodenum, fat, bone marrow, esophagus, pancreas, lung, heart, spleen, blood, brain, thymus, 

thyroid, adrenal, testes, untreated skin, and carcass. 

3.3. METABOLISM 

Despite extensive research, the metabolism of benzene is still not thoroughly understood. 

It is generally accepted that benzene itself is not directly responsible for causing the toxic 

effects; however, the metabolic product or products responsible for the noncancer and 

carcinogenic effects of benzene exposure have not been clearly defined. The evidence suggests 

that several metabolites, as well as interactions between these metabolites, may be necessary to 

explain the toxic effects of benzene. A complete review of the metabolism of benzene is beyond 

the scope of this toxicological review. The reader is referred to several excellent reviews for 

more complete coverage (ATSDR, 1997; Snyder et al., 1993b; Snyder and Hedli, 1996; Ross, 

9 
1996; Witz et al.,1996). This section provides an overview of benzene metabolism, focusing on 

the issues that are relevant to the interpretation of the noncancer effects due to benzene exposure. 

3.3.1. Metabolic Pathways 

The first step in benzene metabolism is the formation of the epoxide, benzene oxide, 

catalyzed by cytochrome P450 2E1 (CYP2E1) (Figure 1). After formation of benzene oxide, the 

metabolism of benzene branches into several alternative metabolic pathways (Jerina et al., 1968; 

Lovern et al., 1997). However, although Jerina et al. (1968) provided a kinetically plausible 

argument for the existence of benzene oxide, they did not show that it was formed from benzene. 

Benzene oxide rearranges nonenzymatically to form phenol, the major product of initial benzene 

metabolism. Alternatively, benzene oxide may react with glutathione (GSH) to form phenylmercapturic acid; undergo enzymatic conversion by epoxide hydrolase to benzene dihydrodiol 

with subsequent formation of catechol; or undergo an iron-catalyzed, ring-opening reaction to 

form trans,trans-muconaldehyde (MUC) with subsequent metabolism to trans,trans-muconic 

acid (MA). Phenol is further oxidized by CYP2E1 catalysis to hydroquinone. Further oxidation 

of hydroquinone to p-benzoquinone is catalyzed by myeloperoxidase (MPO) (Smith et al., 1989). 

All of the phenolic products may be conjugated with sulfate or glucuronic acid, and the 

conjugates of phenol and hydroquinone are the major benzene metabolites excreted in urine 

(Sabourin et al., 1989; Wells and Nerland, 1991). 

Mathews et al. (1998) studied the metabolism of 14C-benzene in male F344 rats over a 

wide range of oral (gavage) doses, 0.02, 0.1, 0.5, or 100 mg benzene/kg body weight, in male 

B6C3F1 mice at oral doses of 0.1 and 100 mg/kg and in male hamsters at 0.02, 0.1 and 100 

mg/kg. In F344 rats, at lower doses (0.02, 0.1, and 0.5 mg/kg), greater than 95% of the dose was 

recovered in the urine within 48 hours, and a small percentage (about 3%) was recovered in the 

breath. At higher doses, 10 and 100 mg/kg, the percentage eliminated in the breath increased to 

about 9 and 50%, respectively. Excretion in the feces was a minor route at all doses. A similar 

pattern of disposition of the radiolabel dose was also observed in mice and hamsters. Mathews 

et al. (1998) also examined the profile of urinary metabolites formed. Interestingly, the 

percentage of prephenylmercapturic acid and phenylmercapturic acid, indicators of benzene 

oxide production, was relatively constant across all doses for rats (~13%), mice (~5%), and 

hamsters (~7%). However, the percentage of hydroquinone and related conjugates ranged from 

about 3% at the highest dose to as much as 7% at the lowest doses. A higher percentage of 

hydroquinone metabolites was seen in mice (~30%) and in hamsters (~30%), but it did not 

appear to be dose dependent. 

Figure 1. Metabolic pathways for benzene. 

Source: Ross, 2000. 

3.3.2. Requirement for CYP2E1 

Oxidation of benzene by the CYP2E1 isoenzyme has been demonstrated to be required 

for the expression of hematotoxicity and genotoxicity. Administration of toluene, a competitive 

inhibitor of benzene metabolism, causes a decrease in benzene metabolite formation and a 

reduction in toxicity (Andrews et al., 1977). The primary oxidation of benzene by CYP2E1 

10 
occurs in the liver, and further metabolism to the final toxic compound occurs in the target 

tissues. Sammett et al. (1979) showed that partial hepatectomy of rats diminished both the rate 

of metabolism of benzene and its toxicity, suggesting that a metabolite and/or metabolites 

formed in the liver are necessary for toxicity. Immunoinhibition studies in rat and rabbit hepatic 

microsomes also have implicated CYP2E1 as the major oxidative isoenzyme involved in 

benzene metabolism (Johansson and Ingelman-Sundberg, 1988; Koop and Laethem, 1992). 

Convincing evidence that CYP2E1 activity is required for expression of benzene toxicity was 

provided using transgenic CYP2E1 knockout mice that do not express hepatic CYP2E1 activity 

(Valentine et al., 1996). Benzene exposure at 200 ppm (640 mg/m3

) for 6 hours/day for 5 days 

resulted in severe genotoxicity and cytotoxicity in wild-type mice, but no toxicity was observed 

in CYP2E1 knockout mice. Thus, the requirement for CYP2E1 metabolism has been clearly 

established, but the identity of the toxic metabolites remains uncertain. 

Recently, Bernauer et al. (2000) investigated a role of CYP2E1 expression in bone 

marrow and its intra- and interspecies variability in rats, rabbits, and humans, because it is a 

target organ for several chemicals, including benzene. The data demonstrated a presence of 

CYP2E1 in the bone marrow of all species investigated, thus supporting the hypothesis of 

CYP2E1-dependent local metabolism of several chemicals that includes benzene possibly 

contributing to myelotoxicity and hematotoxicity. 

11
3.3.3. Toxicity of Benzene Metabolites 

The products of the phenol pathway (catechol, hydroquinone, and p-benzoquinone), MA, 

and benzene epoxide have been proposed as benzene metabolites that may cause the toxic effects 

of benzene exposure. 

3.3.3.1. Phenolic Products 

The production of the initial benzene metabolites occurs primarily in the liver; however, 

the effects of benzene toxicity are primarily expressed as hematotoxicity and myelotoxicity in 

the bone marrow (Snyder and Hedli, 1996). The evidence suggests that secondary metabolism in 

bone marrow is required for expression of the toxicity of benzene (Schlosser and Kalf, 1989; 

Subrahmanyam et al., 1990, 1991). An alternative hypothesis is that metabolism of benzene to 

toxic metabolites can occur in the bone marrow itself. Metabolism of benzene to hydroquinone 

in the bone marrow of rats has been demonstrated, but the amount of metabolites produced was 

small (Irons et al., 1980). The presence of CYP2E1 has been detected in rabbit bone marrow 

(Schnier et al., 1989); however, CYP2E1 could not be detected in the bone marrow of mice, the 

species demonstrated to be most sensitive to benzene hematotoxicity (Genter and Reico, 1994). 

Rickert et al. (1979) observed that catechol and hydroquinone concentrations persisted in the 

bone marrow long after blood levels had declined following inhalation exposure. 

Most research, therefore, has focused on the possibility that phenol, catechol, and 

hydroquinone, generated by reactions in the liver, are transported to the bone marrow and other 

target tissues and subsequently activated by the action of peroxidase (Smith et al., 1989; 

Rushmore et al., 1984; Low et al., 1995). A potential role for conversion of hydroquinone to pbenzoquinone by the peroxidase component of prostaglandin H synthase has also been suggested 

(Schlosser and Kalf, 1989). Subsequent research, however, suggested that the prostaglandin H 

synthase is not involved in benzene toxicity (Ganousis et al., 1992). 

A major problem with the hypothesis that phenolic metabolites are responsible for bone 

marrow toxicity is that administration of phenol fails to duplicate the effects of benzene (Tunek 

et al., 1981; NCI, 1980). Kenyon et al. (1995) suggested that the distribution of phenolconjugating enzymes and benzene-oxidizing enzymes within the liver might have accounted for 

this result. The authors observed that phenol-conjugating enzymes were more concentrated in 

the periportal area of the liver, the first region to absorb orally administered phenol, whereas 

oxidizing enzymes were more concentrated in the pericentral region of the liver. This could 

have lead to rapid excretion of orally administered phenol before it was further metabolized to 

hydroquinone by CYP2E1 in the pericentral region of the liver. A combination of phenol and 

hydroquinone is needed to cause bone marrow toxicity (Eastmond et al., 1987). Intraperitoneal 

(i.p.) injection of either phenol or hydroquinone alone in B6C3F1 mice failed to cause significant 

bone marrow toxicity; however, co-administration of these metabolites caused a reduction in 

bone marrow cellularity with a clear dose-response curve. The presence of phenol apparently 

stimulated peroxidase-dependent metabolism of hydroquinone to p-benzoquinone. Increased 

covalent binding of 14C-hydroquinone was observed in bone marrow when a combination of 

phenol and hydroquinone was administered to mice (Subrahmanyam et al., 1990, 1991). 

12
In contrast to the observations of Eastmond et al. (1987), catechol was reported to 

markedly stimulate peroxidase activation of hydroquinone in murine stroma (Ganousis et al., 

1992). In addition to the effects of phenol on the activity of MPO, co-administration of phenol 

was shown to increase the concentration of phenol and hydroquinone in the blood in B6C3F1 

mice. Legathe et al. (1994) measured the area under the curve (AUC) blood concentration-time 

curve for phenol and hydroquinone administered alone or in combination at the same dose levels 

used by Eastmond et al. (1987). Co-administration increased the phenol AUC by 1.4-fold and 

the hydroquinone AUC by 2.6-fold in comparison to each compound administered alone. 

Legathe et al. (1994) suggested that this resulted from saturation of the enzymes that form sulfate 

and glucuronide conjugates of phenolics in the liver. These results suggest that interactions of 

two or more phenolic metabolites of benzene may be necessary to cause the observed bone 

marrow toxicity. 

3.3.3.2. Trans,trans-Muconaldehyde 

Muconaldehyde (MUC), a highly reactive six-carbon diene dialdehyde, has also been 

proposed to be a benzene metabolite responsible for bone marrow toxicity. Evidence supporting 

this hypothesis has been summarized by Witz et al. (1996). MUC has been shown to cause 

hematotoxicity following short-term exposures. Administration of 2 mg/kg/day MUC to CD-1 

mice for 16 days caused significant decreases in bone marrow cellularity, lymphocytes, red 

blood cell (RBC) counts, hematocrit (HCT) (volume percentage of erythrocytes in whole blood), 

and hemoglobin (Hgb) and significant increases in white blood cell (WBC) count and spleen 

weight (Witz et al., 1985). Snyder et al. (1989) also found that administration of MUC caused 

bone marrow toxicity in mice and that co-administration of MUC and hydroquinone resulted in a 

dramatic decrease in 59Fe incorporation into red cell Hgb. 

Production of MUC is expected as a step in the ring-opening pathway leading to MA. 

Excretion of MA in the urine was demonstrated in rabbits in the early work of Parke and 

Williams (1953), and excretion of MA was also demonstrated in mice, rats, cynomolgus 

monkeys, chimpanzees, and humans (Gad-El Karim et al., 1985; Sabourin et al., 1988a, 1989, 

1992). Metabolism of MUC to MA has been shown in vivo in mice (Witz et al., 1990a,b). 

Urinary excretion of MA has been used as a sensitive biomarker of benzene exposure in humans. 

At low doses, urinary MA concentration was found to be linearly correlated with time-weighted 

average (TWA) benzene exposure concentrations (Bechtold and Henderson, 1993; Bechtold et 

al., 1991). Thus, the ring-opening pathway is active in several animal species. There is no direct 

evidence, however, for the in vivo formation of MUC from benzene. 

Although MUC formation has not been demonstrated in animals in vivo, formation of 

MUC from benzene has been demonstrated in a mouse hepatic microsomal system (Latriano et 

al., 1986; Zhang et al., 1995a). MUC may be too reactive to be isolated from in vivo systems. 

Using isolated rat livers perfused with 0.7 mM benzene solutions through the portal vein, Grotz 

et al. (1994) demonstrated that the complete pathway for formation of MA from benzene was 

active in the liver, but MUC was not detected in the perfusate collected via the hepatic vein. 

When MUC was added to the perfusion solution, it was rapidly and efficiently metabolized to 

MA, and only traces of MUC were detected in the perfusate collected after a single pass through 

the isolated rat livers. Thus, it seems unlikely that sufficient quantities of MUC could reach the 

target tissues by circulation in the blood. The acid-alcohol 6-hydroxy-trans,trans-2,4-

13 
hexadienoic acid (COOH-M-OH) and MA were the major MUC metabolites detected in the rat 

liver perfusate, but both the acid-aldehyde 6-oxo-trans,trans-2,4-hexadienoic acid (COOH-MCHO) and the aldehyde-alcohol 6-hydroxy-trans,trans-2,4-hexadienal (CHO-M-OH) were 

detected as minor peaks (Grotz et al., 1994). The CHO-M-OH has been demonstrated to react 

with GSH, to be cytotoxic to isolated rat hepatocytes, and to be hematotoxic in mice (Goon et 

al., 1993; Zhang et al., 1995b). Thus, CHO-M-OH may be the ring-opened metabolite that 

causes the hematoxic effects of administered MUC, and it may play a role in causing the toxic 

effects of benzene exposure. 

3.3.3.3. Benzene Oxide 

Benzene oxide was believed to be too reactive to escape the liver and to cause toxicity in 

the bone marrow. However, Lovern et al. (1997) recently showed that benzene oxide constituted 

7% of the benzene metabolites after 18 minutes’ incubation with liver microsomes. Lindstrom et 

al. (1997) demonstrated the presence of benzene oxide in the blood and estimated its half-life to 

be about 8 minutes. Using a PBPK model, Lindstrom et al. (1997) predicted that the dose to the 

body from benzene oxide would be about 22-fold greater than from 1,4-benzoquinone. Thus, 

circulating benzene oxide can contribute to observed DNA and protein adduct formation. 

Further research is needed to establish the role of benzene oxide in causing bone marrow 

toxicity. 

3.3.4. Species, Route, and Rate Differences 

Differences in the rates of benzene metabolism and the metabolites formed correlate with 

observed differences in sensitivity to benzene toxicity. Mice are more sensitive than rats to the 

toxic effects of benzene (Huff et al., 1989; Snyder et al., 1978; Ward et al., 1985). Sabourin et 

al. (1987) found that metabolism of benzene, determined by excretion of water-soluble 

metabolites in the urine, became saturated at lower doses in B6C3F1 mice than in F344/N rats. 

The amount of metabolites per kilogram body weight was similar in the two species at gavage 

doses up to 50 mg/kg, but above this level, total metabolites in rats continued to increase but 

total metabolites in mice did not increase further. Following inhalation exposures, total 

metabolites excreted were higher for mice than for rats at all benzene concentrations due to a 

higher amount inhaled by mice. Total metabolite excretion was exponentially related to benzene 

concentration, but the concentration needed to reach half of the maximal metabolite formation 

was lower in mice (220 mg/m3

 [69 ppm]) than in rats (260 mg/m3 [81.5 ppm]). By both oral and 

inhalation routes, the metabolism of benzene was saturated at lower concentrations in mice than 

in rats. Thus, differences in total metabolite formation between rats and mice did not explain the 

greater sensitivity of mice to benzene. 

Sabourin et al. (1988a) compared the metabolites of benzene in the blood, liver, lung, and 

bone marrow of male B6C3F1 mice and F344/N rats following 6 hours of inhalation exposure to 

benzene at 50 ppm (160 mg/m3

). Hydroquinone glucuronide, hydroquinone, and MA were 

present in much higher concentrations in the mouse than in the rat tissues. Thus, metabolism in 

mice leads to much greater exposures to potentially toxic benzene metabolites, which may 

explain the greater sensitivity of mice to benzene toxicity. 

14
Exposure concentration and routes also affect the distribution of metabolites formed in 

mice and rats (Sabourin et al., 1989). Male B6C3F1 mice and F344/N rats were given gavage 

doses of 1, 10, or 200 mg/kg benzene or they were exposed to 5, 50, or 600 ppm (16, 160, or 

1916 mg/m3

) for 6 hours. Water-soluble metabolites were determined in blood, urine, liver, 

lung, and bone marrow. A shift in the distribution from potentially toxic metabolites, 

hydroquinone glucuronide, and MA to the detoxification metabolites, phenylglucuronide and 

prephenylmercapturic acid, was observed from low concentrations to high concentrations 

following both oral and inhalation exposures in mice. A similar shift was observed in rats, 

except that hydroquinone glucuronide was a minor metabolite in rats at all concentrations. Thus, 

extrapolation from toxicological studies conducted at high exposure concentrations may 

significantly underestimate risks at lower exposure concentrations. There was also no simple 

correlation between the dose of metabolites to tissues resulting from oral and inhalation 

exposures. This may be due to saturation of benzene metabolism following bolus doses of 

benzene administered by gavage. 

Urinary excretion of benzene metabolites has also been examined in cynomolgus 

monkeys and chimpanzees (Sabourin et al., 1992). The proportion of an oral gavage dose 

excreted in the urine decreased from 50 to 15% as the dose rate increased from 5 to 500 mg/kg in 

cynomolgus monkeys. Phenyl sulfate was the primary metabolite excreted, and the proportion 

excreted as hydroquinone conjugates and MA decreased as the dose increased, as was observed 

previously in mice and rats. The proportion excreted as MA decreased from 4.4% at the low 

dose to 1.3% at the high dose. At all levels the proportion of MA excreted as MA was much 

lower in the monkey than in either mice or rats. Chimpanzees were exposed at only one low 

dose rate of 1 mg/kg by intravenous injection. Less than 15% of the administered dose was 

recovered as hydroquinone conjugates and MA, and 79% was recovered as phenyl conjugates. 

On the basis of the urinary metabolite profiles, the mouse appears to metabolize the largest 

fraction of benzene via pathways leading to hydroquinone conjugates and MA (67% at low 

doses) followed by monkeys (31%), rats (17%), and chimpanzees (14%). Thus, there are 

apparently large quantitative differences among species in the metabolism of benzene to 

potentially toxic metabolites. Because few data are available on the proportion of benzene 

metabolized to potentially toxic metabolites in humans, considerable uncertainty exists in 

determining which animal model best represents human metabolism. 

Henderson (1996) reviewed the species differences in benzene metabolism. The 

pathways of benzene metabolism appear to be similar in all species studied; however, there are 

quantitative differences in the fraction of benzene metabolized by different pathways. Monkeys 

and mice metabolize more of the benzene dose to hydroquinone metabolites than do rats or 

chimpanzees, especially at low doses. Mice appear to have a greater overall capacity to 

metabolize benzene than do rats and primates. This finding may explain why mice are more 

sensitive than rats to benzene. In all species, a greater proportion of benzene is converted to 

hydroquinone and ring-open metabolites at low doses than at high doses. 

3.3.5. Induction of CYP2E1 

Benzene exposure has been found to induce CYP2E1 activity, thereby increasing the rate 

of toxic metabolite formation. Pretreatment of mice, rats, and rabbits subcutaneously with 

benzene increased benzene metabolism in vitro without increasing total cytochrome P450 

15 
concentrations (Arinc et al., 1991; Gonasun et al., 1973; Saito et al., 1973). In contrast, Sabourin 

et al. (1990) found no significant effect on the metabolism of benzene when F344 rats and 

B6C3F1 mice were pretreated by inhalation exposure to 600 ppm (316 mg/m3

) of benzene. The 

rate of benzene metabolism can be altered by pretreatment with various compounds. CYP2E1 

also metabolizes alcohol and aniline, and CYP2E1 can be induced by these substrates (Chepiga 

et al., 1990; Parke, 1989; Snyder et al., 1993a). Phenol, hydroquinone, benzoquinone, and 

catechol have also been shown to induce P450 in human hematopoietic stem cells (Henschler 

and Glatt, 1995). Therefore, exposure to chemicals that stimulate the activity of this enzyme 

system prior to exposure to benzene could increase the rate of benzene metabolism. 

Daiker et al. (1996) found that repeated oral benzene exposure of female B6C3F1 mice 

for 3 weeks at 50 mg/kg/day decreased CYP2E1 activity by 34% and activated the detoxification 

enzyme GSH transferase by 30% without affecting aldehyde dehydrogenase, another detoxifying 

enzyme. The authors suggested that these changes in enzyme activity may serve a protective 

role against repeated benzene exposure. 

Exposure to benzene was found to interfere with the disappearance of ethanol from the 

body in rats (Nakajima et al., 1985). The results of further studies showed that ethanol treatment 

increased the production of hydroxylated benzene metabolites, phenol, and hydroquinone, 

suggesting induction of benzene metabolism (Nakajima et al., 1987). The possibility therefore 

exists of a synergism between alcohol and benzene. 

3.3.6. Mechanism of Toxicity 

Just as several benzene metabolites have been implicated in benzene toxicity, several 

different mechanisms also may contribute to the overall toxic effects of benzene, as discussed 

below. 

3.3.6.1. Formation of Covalent Adducts 

Benzene metabolites form covalent adducts with both cell proteins and DNA; however, 

the role of adduct formation in toxicity is unclear. Treatments that reduce benzene toxicity also 

reduce adduct formation. Partial hepatectomy of rats resulted in reduced hematotoxicity and also 

in reduced covalent binding in bone marrow (Sammett et al., 1979). Levels of adduct formation 

in the hematopoietic tissues of mice strains have been shown to correlate with the relative 

sensitivity of the strains (Longacre et al., 1981). 

Covalent binding of benzene metabolites to DNA was first demonstrated by Lutz and 

Schlatter (1977). Mazzullo et al. (1989) found that formation of DNA adducts of benzene 

metabolites was linear at low benzene concentrations, but saturation of adduct formation 

occurred at high benzene concentrations. Hedli et al. (1996) investigated DNA adduct formation 

from the benzene metabolites hydroquinone and 1,2,4-benzenetriol in combination with 

investigations of the effects of these two metabolites on cell differentiation in a hematopoiesis 

model system. Hydroquinone formed DNA adducts in human promyelocytic leukemia cells, but 

1,2,4-benzenetriol did not. Both metabolites, however, inhibited retinoic acid-induced 

maturation of human promyelocytic leukemia cells to granulocytes. Thus, DNA adduct 

formation may be important in hydroquinone but not in 1,2,4-benzenetriol toxicity. 

16 
Creek et al. (1997) showed that DNA and protein adduct formation following i.p. 

administration of 14C-benzene was linear over a dose range spanning eight orders of magnitude 

in B6C3F1 mice. At doses greater than 16 mg/kg body weight, however, adduct formation was 

nonlinear. This corresponds very closely to the metabolic saturation level of 15 mg/kg body 

weight observed by Sabourin et al. (1987). Benzene was administered to male B6C3F1 mice 

over a dose range of 700 pg/kg to 500 mg/kg body weight. Liver DNA adduct levels peaked at 

0.5 hours after exposure, but bone marrow DNA adduct levels peaked between 12 and 24 hours. 

These results indicate that adduct formation is linear in the range of benzene concentrations 

where benzene toxicity is first detected and becomes saturated at higher doses where most 

toxicity experiments have been conducted. 

Although the mechanism by which adduct formation results in bone marrow toxicity is 

not well established, a few experiments have suggested that binding of hydroquinone or pbenzoquinone to sulfhydryl groups at the active sites of proteins could be responsible for the 

genotoxic effects of benzene. Hydroquinone and p-benzoquinone interfere with binding of 

guanosine triphosphate (GTP) to tubulin by alkylating nucleophilic sulfhydryl groups (Irons and 

Neptun, 1980; Pfeifer and Irons, 1983). Binding of GTP is required for stabilization of tubulin 

polymerization during microtubule formation and, therefore, interferes with spindle formation 

during mitosis. Several metabolites of benzene have also been shown to inhibit the activity to 

human topoisomerase II (Frantz et al., 1996). Topoisomerase II enzymes relieve torsional strain 

on DNA during replication and transcription and are also believed to function during 

recombination and chromosome condensation. Both p-benzoquinone and MUC directly 

inhibited topoisomerase at concentrations as low as 10 M. With the addition of horseradish 

peroxidase and peroxide to the assay mixture, the benzene metabolites phenol, 4,4'-biphenol, 

2,2'-biphenol, hydroquinone, catechol, and 1,2,4-benzenetriol all inhibited topoisomerase II. 

Topoisomerase inhibitors tend to be strong clastogens in mammalian cells. Thus, inhibition of 

topoisomerase II could lead to the clastogenic effects observed following benzene exposure. 

Addition of GSH to the assay mixture also protects against topoisomerase II inhibition, thus 

suggesting that interaction of benzene metabolites with an essential sufhydryl residue occurs. 

Hutt and Kalf (1996) also reported that p-benzoquinone inhibits the activity of topoisomerase II. 

Protein adduct formation by benzene metabolites differs greatly between mice and rats. 

McDonald et al. (1994) used uniform 14C-benzene to determine total protein binding and C6- 13C-benzene to determine formation of adducts due to the reactions of benzene oxide, 1,4-

benzoquinone, or 1,2-benzoquinone. Total binding to protein and specific binding to cysteine 

residues were estimated for blood Hgb and bone marrow total protein. Formation of adducts of 

benzene oxide, 1,4-benzoquinone, and 1,2-benzoquinone accounted for 74% of the binding to 

cysteine residues in rat Hgb but for only approximately 25% of the cysteine binding in mouse 

Hgb or bone marrow proteins of either species. Thus, other benzene metabolites must also bind 

to cysteine. Adducts of benzene oxide were highest in rat Hgb but accounted for only a small 

proportion of adduct formation in mouse Hgb. Benzene oxide adducts accounted for less than 

3% of total adduct formation in the bone marrow of either mice or rats, and benzoquinone adduct 

formation accounted for a much greater proportion of total protein binding in bone marrow. 

However, 1,2-benzoquinone adducts were more prevalent in rat bone marrow proteins and 1,4-

benzoquinone adducts were more prevalent in mouse bone marrow proteins. This suggests that 

there are significant differences in benzene metabolism between mice and rats. A background 

level of both 1,2- and 1,4-benzoquinone protein adducts that greatly exceeded the level of 

17 
labeled adduct formation was also observed in bone marrow proteins of control mice and rats. 

This result suggests that low doses of benzene may not cause toxicity through protein adduct 

formation. 

3.3.6.2. Genotoxicity 

The association between benzene exposure and the appearance of structural and 

numerical chromosomal aberrations in human lymphocytes suggests that benzene may be 

considered as a human clastogen. In in vivo animal studies, benzene induced cytogenetic 

effects, including chromosome and chromatid aberrations, sister chromatid exchanges, and 

micronuclei (Anderson and Richardson, 1981; Au et al., 1991; Erexson et al., 1986; Fujie et al., 

1992; Kolachana et al., 1993; Siou et al., 1981; Toft et al., 1982; Ward et al., 1992). Several 

lines of evidence also indicate that benzene is genotoxic in humans under occupational exposure 

conditions (Ding et al., 1983; Sasiadek et al., 1989; Yardley-Jones et al., 1990; Major et al., 

1992; Eastmond, 1993; Tompa et al., 1994). However, these studies lacked good exposure 

monitoring data, involved multiple chemical exposures, and were often poorly designed, with 

inappropriate control groups. 

Benzene has been shown to produce DNA breaks in Chinese hamster ovary cells 

independent of metabolic activators; however, in vitro assays indicate that genotoxicity of 

benzene is primarily due to its metabolites (Zhang et al., 1993; Eastmond et al., 1994). Benzene 

is known to affect cell cycle progression, RNA and DNA synthesis, and DNA binding (Forni and 

Moreo, 1967). Chen and Eastmond (1995) showed that benzene metabolites can adversely affect 

human topoisomerases; however, some DNA repairs may occur in human cells (Chenna et al., 

1995). 

Even with these studies, no data exist on the quantitative relationship between measured 

benzene exposures and clastogenic effects. 

Benzene itself is not mutagenic in short-term assays in either bacterial or animal systems 

(Dean, 1985). Mutagenicity of benzene metabolites, however, is well established. 

Hydroquinone, catechol, 1,2,4-trihydroxybenzene, and trans-1,2-dihydrodiol were active in 

inducing elevated resistance to 6-thioguanine in Chinese hamster V79 cells. MUC is an active 

mutagen in V79 cells, but it is only weakly active in Salmonella (Witz et al., 1990a). 

3.3.6.3. Oxidative Stress 

Benzene may also produce oxidative stress in the target tissues. The benzene metabolite 

p-benzoquinone is highly reactive and can deplete cellular levels of GSH (Brunmark and 

Cadenas, 1988). Benzene metabolites can also be involved in redox cycling, resulting in the 

production of reactive oxygen species that can also react with macromolecular components (Rao 

and Snyder, 1995). 

18
3.3.6.4. Inhibition of Cytokine Formation 

The stromal microenvironment of the bone marrow that normally modulates stem cell 

proliferation and differentiation is a potential target for the hematotoxicity of benzene (Cox, 

1991; Snyder et al., 1989; Kalf, 1987). The interaction of the stroma with the stem cells is 

necessary for hematopoiesis. Furthermore, the stromal macrophage produces interleukin-1 (IL1), a cytokine also essential for hematopoiesis. Patients with aplastic anemia usually exhibit 

monocyte dysfunction and decreased IL-1 production (Renz and Kalf, 1991). 

Renz and Kalf (1991) demonstrated the disruption of IL-1 production by the stromal 

macrophages of mice exposed to benzene. The mice were injected i.p. with 600 or 800 

mg/kg/day benzene for 2 days. The stromal macrophages removed from these mice and cultured 

with lipopolysaccharide produced the IL-1 precursor 34-kilodalton (Kd) pre-interleukin-1alpha 

(IL-1) but could not convert the precursor to the 17-Kd mature cytokine. Hydroquinone added 

in vitro also inhibited the conversion of the pre-IL-1 to the mature cytokine in mouse 

macrophages. However, administration of recombinant mouse IL-1 to mice before a bonemarrow-suppressing dose of benzene ameliorated the bone marrow depression, probably by 

circumventing the inability of the stromal macrophage in benzene-treated animals to process preIL-1 to the mature cytokine. Thus, Renz and Kalf (1991) suggested that benzene-induced 

depression of bone marrow cellularity may result from the failure of the stromal macrophages to 

process pre-IL-1 to mature IL-1, which activates the stromal fibroblast production of the 

colony-stimulating factor required for the differentiation of stem cells. 

Niculescu et al. (1995, 1996) demonstrated that p-benzoquinone, the oxidation product of 

hydroquinone in the cell, causes a concentration-dependent inhibition of highly purified human 

platelet calpain with an IC50 of 3 M. Calpain is a calcium-activated, cysteine-dependent 

protease that catalyzes the processing of pre-IL-1 to the mature cytokine in vivo. The 

investigators also demonstrated that p-benzoquinone inhibits the processing of interleukin-1

(IL-1), the product of a distinct second gene, to IL-1 by the sulfhydryl-dependent protease 

referred to as IL-1 convertase. 

3.4. ELIMINATION AND EXCRETION 

3.4.1. Oral Exposure 

The information base on the clearance of benzene from the body after oral exposure to 

the compound is limited to reports of studies in experimental animals. For example, in one of a 

series of toxicological reports on the metabolism and excretion of radiolabeled xenobiotics, 

Parke and Williams (1953) recovered a substantial proportion of the administered dose on the 

breath as unmetabolized product (43%). This amount compared with 33% urinary excretion in 

which the label partitioned into such components as conjugated phenols (23% of the dose), 

hydroquinone (4.8%), and catechol (2.2%), among others. Residual amounts of radioactivity 

remained deposited in the tissues or were excreted in the feces. 

In a study using male C57BL/6N mice given single oral doses of 14C-benzene (10 or 200 

mg/kg), McMahon and Birnbaum (1991) reported the effects of age on benzene metabolism. 

Radioactivity was monitored in urine, feces, and breath. Consistent with previous reports, the 

19 
following urinary metabolites were detected: hydroquinone glucuronide, MA, phenyl 

glucuronide, phenyl sulfate, catechol glucuronide, hydroquinone sulfate, and prephenylmercapturic acid. At various time points up to 48 hours after dosing with 10 mg/kg, a 

significant decrease in the urinary excretion of benzene-derived 14C was observed for 18- versus 

3-month-old mice. When expressed as the percent of total benzene administered, the relative 

amounts of some urinary metabolites varied between the 10 and 200 mg/kg dose groups, thereby 

indicating dose-related quantitative changes in the urinary excretion of benzene metabolites. At 

the low dose (10 mg/kg), urinary excretion was the major route of elimination. Hydroquinone 

glucuronide, phenyl sulfate, and MA were the major metabolites at this dose, accounting for 

40%, 28%, and 15% of the dose, respectively. At 200 mg/kg, urinary excretion decreased to 

account for 42–47% of the administered dose, whereas respiratory excretion of volatile 

components increased to 46–56% of the administered dose. Fecal elimination was minor and 

relatively constant over both doses, accounting for 0.5–3% of the dose. Although age-related 

differences in benzene disposition were observed, they could be attributed to alterations in 

physiological function occurring with age; the significance of toxicity versus aging could not be 

ascertained. 

Fluctuation in the proportion of radiolabeled benzene excreted in urine was observed 

with increasing dose by Sabourin et al. (1987), who exposed B6C3F1 mice, F344 rats, and 

Sprague-Dawley rats via gavage to doses ranging from 0.5 to 300 mg/kg 14C-benzene. At the 

lower doses (< 15 mg/kg), near quantitative amounts of label were recovered in various urinary 

metabolites, whereas at the higher doses, unchanged 14C-benzene began to appear in expired air 

in increasing proportions. For example, at doses of 150 mg/kg and above, greater than 50% of 

the load was cleared on the breath as unchanged starting compound. These data point to the 

saturation of the metabolic processes with increasing dose, although even at lower doses a 

degree of fluctuation in the metabolic pattern was implied by dose-dependent changes in the 

profile of metabolic products detected in the urine. 

3.4.2. Inhalation Exposure 

In a key report of studies that addressed the metabolic fate of inhaled benzene issue in 

humans, Nomiyama and Nomiyama (1974) showed that at least a proportion of the absorbed 

compound can be excreted in the urine as sulfate- or glucuronide-conjugated phenols or MAs. 

At the benzene concentrations employed (52–62 ppm [166–198 mg/m3

]), respiratory uptake was 

considered to be in the region of 47% and respiratory excretion for the 4-hour exposure period 

was approximately 17%. These values were in reasonably good agreement to those obtained in 

an earlier study by Srbova et al. (1950), who showed that the respiratory excretion of retained 

benzene could vary between 16.4 and 41.6% across a 7-hour exposure period. 

In general, insufficient data exist to unequivocally assign one elimination route or 

another as being of primary importance when human beings are exposed to benzene via 

inhalation. For example, Sherwood (1988), using a complex experimental protocol, employed a 

single human subject who was alternately exposed to either 6.4 ppm (20 mg/m3

) benzene for 8 

hours or 99 ppm (316 mg/m3) for 1 hour to monitor the kinetics of benzene elimination via the 

various routes. Sherwood separated the release of benzene on the breath into several distinct 

phases and was able to show that a greater proportion of the total dose was excreted in urine 

rather than via expiration. These results also showed that urinary excretion of phenol conjugate 

20 
was biphasic, with an initial rapid excretion phase followed by a slower excretion phase. The 

importance of urinary excretion was also emphasized by the work of Inoue et al. (1986), who 

showed a good correlation between urinary phenol levels and benzene exposure across a 

relatively wide concentration range (1–200 ppm [3.2–640 mg/m3

]). 

Occupational exposure and cross-sectional studies have also pointed to the appearance of 

benzene metabolites in urine as a consequence of exposure to such agents as sidestream tobacco 

smoke (Bartczak et al., 1994) and gasoline vapors (Lagorio et al., 1994) or when the urine 

content of smokers versus nonsmokers was compared (Kok and Ong, 1994; Melikian et al., 

1994). In general, one cigarette delivers an average of 55 μg of benzene, and an average smoker 

takes in about 1.2–1.8 mg of benzene per day from smoking. Popp et al. (1994) detected 

muconic acid and S-phenyl-N-acetyl cysteine levels in the urine of car mechanics at levels that 

correlated with levels of the compound in the bloodstream and the breathing zone. These 

observations support the earlier suggestion by Ghittori et al. (1993) that benzene and its 

metabolites in the urine may be an important biomarker of occupational exposure to the 

compound. 

An accumulation of experimental data in laboratory animals has shown an essentially 

similar pattern of benzene elimination and excretion as in human beings. In broad terms, this is 

characterized by the release of unchanged compound on the breath and the appearance of 

metabolites in the urine. For example, at high incidental concentrations of benzene (500 ppm 

[1597 mg/m3

] for 6 hours), an initial rapid phase of elimination was followed by a slower phase 

of much longer duration (up to 13 hours or more) (Rickert et al., 1979). 

Perhaps the most rigorous experiments that have examined the dose-response effects of 

benzene via inhalation were those carried out by Sabourin et al. (1987) in parallel to those 

already described for the oral route. The experimental protocol featured a 6-hour, nose-only 

exposure to concentrations of 14C-benzene ranging from 10 to 1000 ppm (32 to 3195 mg/m3). At 

the lower concentrations, less than 6% of the radioactivity was expired on the breath. However, 

at the higher concentrations of benzene (> 850 ppm [2718 mg/m3

]), both rats and mice exhaled 

considerable proportions of the unchanged compound, amounting to 48% of the load in rats and 

14% in mice. In a similar experiment, Dow Chemical Co. (1992a) reported that Duroc-Jersey 

pigs exposed to 0, 20, 100, or 500 ppm (0, 64, 319, or 1597 mg/m3) benzene for 6 hours/day, 5 

days/week for 3 weeks excreted levels of phenol in the urine that increased linearly according to 

dose. 

3.4.3. Dermal Exposure 

Only limited data on excretion of benzene after dermal exposure in humans were found. 

Franz (1984) reported that four volunteers exposed to 14C-benzene (0.0024 mg/cm2

) on the skin 

excreted trace amounts of label in the urine over a 36-hour period, suggesting the ability of the 

compound to penetrate the dermal barrier. Similar results were obtained in monkeys and 

minipigs, with excretion of the compound in monkeys approaching 0.065% (range 

0.033–0.135%) of the applied dose, compared with 0.042% (range 0.030–0.054%) in minipigs 

(Franz, 1984). 

3.4.4. Other Routes of Exposure 

21 
This section discusses very limited data available from three studies in which benzene 

was either given intraperitoneally or subcutaneously. In these studies, the exposure was limited 

to either a single injection or daily injections for up to 3 days. 

Experiments by Cornish and Ryan (1965) used the i.p. route of administration to examine 

the influence of the fed or fasted state on the rates of production of urinary metabolites such as 

conjugated or unconjugated phenols, components that appeared to be increased markedly in the 

fasted state versus fed. In line with their findings of the dose-response effects of benzene when 

administered via gavage or inhalation, Sabourin et al. (1987) found that the proportion of 

radioactivity (mainly as phenyl sulfate) excreted in the urine decreased with increasing dose 

when monkeys were injected intraperitoneally with concentrations of 14C-benzene ranging from 

5 to 500 mg/kg, again suggesting that benzene metabolism can be saturated at the higher dose 

levels. Finally, Longacre et al. (1981) exposed C57BL/6N and DBA/2 mice to benzene 

subcutaneously and monitored the appearance of conjugated phenol, catechol, and hydroquinone 

in the urine, although with varying proportions when one strain was compared with another. 

Amounts of phenyl mercapturic acid were similar, however. 

3.5. PHYSIOLOGICALLY BASED PHARMACOKINETIC MODELS 

PBPK models can be used to address the uncertainty of extrapolating from data of an 

experimental study to a hypothetical administered dose for the same response in other animals or 

humans. The approach seeks to predict the dose-response characteristics of a chemical’s 

potential toxicity through an understanding of the underlying mechanisms by which a substance 

is absorbed, transported to, and metabolized in the primary target organs, on the assumption that 

the biochemical processes involved in the toxicological response will be the same for both 

humans and the animal model(s) in which the harmful effects of the compound had been 

demonstrated. Necessary inputs to the model include a physiologically realistic discrimination 

of a chemical’s movement between metabolically and functionally linked compartments of the 

body and the information necessary to describe this movement mathematically in terms of (1) the 

kinetics of metabolism, (2) rates of movement to and from different target organ groupings, and 

(3) the partitioning of a compound between physiological media. Thus, by modeling a target 

organ-specific internal dose surrogate (parent compound or metabolite, as applicable) from the 

NOAEL or LOAEL of an experimental study, an equivalent allometrically scaled internal dose 

surrogate in human beings can be used to back-extrapolate to a hypothetical “effective” dose for 

the same toxicological response in human beings. 

Potential advantages of the PBPK approach reside in its ability to quantitatively address 

interspecies differences and to take into account the nonlinearity of biological processes when 

extrapolating outside the range of available experimental data. Limitations reside in the 

accuracy/validity of the estimates that constitute the inputs to the model and in the 

oversimplification that is inherent in seeking to define the movement and metabolism of a 

xenobiotic or its metabolites in terms of a small number of functionally defined 

subcompartments. 

Several attempts have been made to develop PBPK models to define the understanding of 

interactions between benzene metabolism and toxicity. The first model for benzene was 

developed by Sato (Sato and Nakajima, 1979; Sato, 1988), who exposed three men to 25 and 100 

22 
ppm (80 and 319 mg/m3

) benzene vapor for 2 hours and then observed a triexponential decay of 

benzene from their blood. The investigators constructed a three-compartment model consisting 

of richly perfused tissues, poorly perfused tissues, and fat, which acted as a major sink for 

benzene. Subsequently, PBPK models have been developed to take into account differences in 

benzene metabolism between species and individuals using both experimental data and 

simulations (Medinsky et al., 1989a; Travis et al., 1990a, b; Bois et al., 1991b; Cox, 1991, 1996). 

A model (Medinsky et al., 1989a, b, c) was developed to describe and predict the fate of 

benzene and to determine whether differences in the metabolic rates between rats and mice could 

explain the differences in toxicity between these species. The model indicated that for inhalation 

concentrations up to 1000 ppm (3195 mg/m3

), mice metabolize at least two to three times as 

much benzene as rats. However, following oral exposure, rats metabolized more benzene on a 

body-weight basis than did mice at doses greater than 50 mg/kg. Patterns of metabolites also 

differed between rats and mice. Mice produced primarily hydroquinone glucuronide and MA 

metabolites linked to toxic effects; on the other hand, rats produced primarily phenyl sulfate, a 

detoxification product. These simulated results agree with experimental data and provide a 

framework for understanding the greater sensitivity of the mouse to benzene toxicity. 

The Medinsky model was based on an earlier PBPK model developed by Ramsey and 

Anderson (1984). The tissue compartments initially included in the model were (1) the liver, 

presumed to be the only organ where benzene metabolism takes place; (2) a group of poorly 

perfused tissues, including muscle and skin; (3) a group of richly perfused tissues, including 

bone marrow, kidney, and intestines; and (4) a fat compartment. All metabolism of benzene, 

consisting of initial metabolism to benzene oxide that is then further metabolized by one of four 

pathways, was modeled by Michaelis-Menten kinetic parameters. Metabolic rate constants were 

determined by fitting the results of model simulations to experimental data obtained by exposing 

mice and rats to benzene orally and by inhalation (Medinsky et al., 1989b; Sabourin et al., 1987). 

However, Bois et al. (1991b) found that the Medinsky model did not simulate the data of Rickert 

et al. (1979) very well, thereby indicating a need for the model to be further refined. 

Since the original Medinsky model was published, additional compartments have been 

added to reflect an advancing understanding of benzene metabolism, and specific biochemical 

and toxicokinetic parameters have been refined to reflect age, sex, and species-specific 

differences (Schlosser et al., 1993; Seaton et al., 1994; McMahon et al., 1994; Kenyon et al., 

1995). Seaton et al. (1994) measured a 13-fold variability in CYP2E1 activity in human hepatic 

microsomes and compared this to the activity in mouse and rat liver microsomes. The model 

predicted the dependence of benzene metabolism on the measured CYP2E1 activity, and the 

proportion of hydroquinone (the suspected toxic metabolite) produced in vitro was correlated 

with the level of CYP2E1 activity. Seaton et al. (1995) measured the initial rates of the two 

major conjugation reactions, phenol sulfonation and hydroquinone glucuronidation, in the 

hepatic microsome preparations of humans, rats, and mice. This information was used in a 

physiological compartment model to predict steady-state concentrations of phenol and 

hydroquinone in blood. 

Among humans, predicted steady-state concentration of phenol varied sixfold (0.38–2.17 

nM), and predicted hydroquinone concentrations varied fivefold (6.66–31.44 nM). Predicted 

steady-state concentrations of phenol were higher in mice than in rats (2.28 vs. 0.83 nM), and 

23 
predicted hydroquinone concentrations were also higher in mice than in rats (42.44 vs. 17.99 

nM). The predicted concentrations for mice were higher than the range for humans, but the rat 

values were within the predicted concentrations for humans. On this basis, the authors suggested 

that the rat may be a good model for humans with respect to tissue dosimetry for these benzene 

metabolites. The authors also suggested that the mouse might be more sensitive than the human 

and that in vitro metabolism data must always be placed within the context of the whole animal’s 

physiology. 

The Medinsky PBPK model has served to organize the available information into a 

coherent model that has helped to refine the specific experimental approaches used to fill the 

gaps in the understanding of the mechanism of benzene toxicity. Although current PBPK 

models may provide insights about putative toxic metabolites and potential biochemical 

mechanisms, they are insufficiently developed to be able to reduce scientific uncertainty 

(Medinsky et al., 1995, 1996; Medinsky, 1995). 

Travis et al. (1990a, b) also developed a model to describe the pharmacokinetics of 

benzene in rats, mice, and humans. The model contains five compartments, consisting of liver, 

fat, bone marrow, muscle, and organs such as brain, heart, kidney, and viscera. The different 

compartments are connected by the arterial and venous blood pathways. Metabolism of benzene 

is assumed to follow Michaelis-Menten kinetics in all species and is assumed to occur primarily 

in the liver and to a lesser extent in the bone marrow. Model simulations were compared with 

experimental data from Sabourin et al. (1987, 1988a), Andrews et al. (1977, 1979), Nomiyama 

and Nomiyama (1974), Snyder et al. (1981), Sato et al. (1975), and Rickert et al. (1979). The 

Travis model successfully simulated uptake, metabolism, and excretion of benzene for mice, 

rats, and humans using experimental data from the studies that were used to develop the model. 

However, the model is of limited value because it does not predict the kinetics of benzene 

metabolites (Bois et al., 1991b). In addition, the concentration of benzene in fat is poorly 

predicted. 

The model developed by Bois and Paxman (1992) provided evidence that exposure rate 

had a strong influence on the rate of formation of several important metabolites of benzene. This 

model has three components describing the pharmacokinetics of benzene and the formation of 

metabolites in the rat. Distribution and elimination of benzene from a five-compartment system 

composed of liver, bone marrow, fat, poorly perfused tissues, and well-perfused tissues make up 

the first component of the model. The bone marrow compartment is included for its relevance to 

human leukemia. Parameter values for this component were derived from the literature and from 

the previously published work of Rickert et al. (1979) and Sabourin et al. (1987). The second 

component describes the metabolic transformations of benzene and its by-products in the liver 

and bone marrow. The reactions are assumed to follow Michaelis-Menten kinetics, except for 

the transformation of benzene oxide into phenol, which occurs spontaneously and may be 

described as a first-order reaction. The third component is the distribution of phenol. In addition 

to the compartments described for benzene, phenol is also assumed to distribute to the lung and 

gastrointestinal tract. 

The model was validated against the data of Cassidy and Houston (1984), Sabourin et al. 

(1987, 1988a, 1989), and Sawahata and Neal (1983). It was also used to predict metabolite 

production for male rats exposed to benzene (nose only) at three different concentrations and for 

24 
three different exposure durations in comparison with the experimental data of Sabourin et al. 

(1989). The three exposure regimens were established to maintain a constant concentration/time 

product. Simulation results indicated that the model may over- or underestimate the level of 

urinary metabolites. 

More recent efforts on development of the Bois and Paxman model have focused on 

defining the physiological pharmacokinetic parameter distributions needed to develop models 

useful in risk assessment (Spear and Bois 1994; Spear et al., 1991; Watanabe and Bois, 1996; 

Bois et al., 1991a, 1996). Spear and Bois (1994) described the outcome of their modeling efforts 

to explain the basis for the paradoxical observation that although phenol is a major initial 

metabolite of benzene, a known carcinogen, a National Cancer Institute chronic study (NCI, 

1980) did not demonstrate carcinogenic activity for phenol. The approach selected was to apply 

Monte Carlo methods using parameter distributions coupled with a pass-fail fit criterion. The 

advantage of this approach is that it acknowledges that in most biological applications, there is 

no clear way to select a “best” set of fixed parameters. On the basis of the researchers’ modeling 

effort, hydroquinone was rejected as the ultimate toxic agent, and the pathway through benzene 

glycol to catechol and MUC appeared to provide a better fit to the data. 

Kenyon et al. (1995) investigated the metabolism of phenol. Even though phenol is 

thought to be a key intermediate in benzene metabolism leading to toxicity, orally administered 

phenol is neither carcinogenic or genotoxic (NCI, 1980). The authors found markedly higher 

excretion of hydroquinone glucuronide after oral benzene exposure as compared with phenol. 

Also, phenol sulfate and phenol glucuronide excretion was much lower following benzene 

exposure than following phenol exposure. This could be explained by differences in the zonal 

distribution of CYP2E1 and detoxification enzymes in the liver. Phenol initially entering the 

liver had a relatively greater chance for conjugation (sulfonation or glucuronidation) in 

periportal hepatocytes of zone 1 than of oxidation by CYP2E1 located in pericentral hepatocytes 

in zone 3. Benzene, on the other hand, was more likely to pass through to zone 3 and be 

oxidized to phenol. 

Watanabe and Bois (1996) examined three methods (multiplicative, additive, and 

allometric) to extrapolate physiological parameter distributions across species, specifically from 

rats to humans. The results indicated that the multiplicative and allometric techniques were able 

to extrapolate the model parameter distributions. 

Bois et al. (1996) applied techniques from population pharmacokinetics, Bayesian 

statistical inference, and physiological modeling to model distribution and metabolism in 

humans. Statistical distributions for the parameters of a physiological model of benzene were 

derived on the basis of existing data. The relationship between the fraction of benzene 

metabolized in bone marrow and benzene exposure was linear up to 10 ppm (32 mg/m3

). The 

median population estimate of the fraction metabolized in bone marrow was 52% (90% 

confidence interval [CI] 47–67%). At levels approaching occupational inhalation exposure 

(continuous 1 ppm [3.2 mg/m3]), the estimated amount metabolized in bone marrow ranged from 

2 to 40 mg/day. However, this model has not been tested for its ability to predict data from other 

studies (Smith and Fanning, 1997). 

25
In summary, PBPK models continue to improve as additional data become available and 

are incorporated and as additional techniques from other scientific fields are applied to modeling 

benzene dosimetry. However, the current published models are insufficiently refined to allow 

them to predict human metabolism accurately (Smith and Fanning, 1997). The key areas for 

refinements appear to be the inclusion of the kinetics of the putative toxic metabolites of benzene 

or their stable precursors. If benzene metabolites such as hydroquinone/ benzoquinone, MUC, 

and/or benzene oxide are the toxic species, then PBPK models need to include descriptions of 

their kinetics if they are to be useful in improving uncertainties in risk assessment. 

3.6. TOXICOKINETICS SUMMARY 

Benzene is almost completely absorbed from the gastrointestinal tract over a wide range 

of dose levels in rats and mice. In contrast, several human studies have indicated that the 

respiratory absorption of benzene is approximately 50%. At a concentration of 10 ppm (31.9 

mg/m3

), retention of benzene during a 6-hour exposure was 33% in rats and 50% in mice. 

Retention decreased at higher concentrations (Sabourin et al., 1987). At high concentrations, 

benzene metabolism is saturated, and respiratory excretion of unmetabolized benzene increases. 

Dermal absorption is less than 1% of the applied dose due to rapid volatilization from 

nonoccluded skin. 

Benzene is rapidly distributed throughout the body regardless of the exposure route. 

Following oral exposure, Low et al. (1989) found the highest concentrations in liver and kidneys, 

intermediate concentrations in blood, and the lowest concentrations in the Zymbal gland, nasal 

cavity, oral cavity, mammary gland, and bone marrow. Benzene is preferentially stored in fat 

(Rickert et al., 1979). Metabolites of benzene are also found throughout the body, with phenol 

being eliminated from bone marrow more rapidly than catechol or hydroquinone (Rickert et al., 

1979). 

Despite extensive research, the metabolism of benzene to toxic metabolites is still not 

completely understood. Metabolism of benzene by CYP2E1 is necessary for the expression of 

hematotoxicity (Sammett et al., 1979; Valentine et al., 1996). Following initial oxidation by 

CYP2E1 to benzene oxide, however, the benzene metabolic pathway branches to produce 

several putative toxic metabolites. Most research has focused on the hypothesis that phenol, 

catechol, and hydroquinone are produced in the liver and transported to the bone marrow, where 

hydroquinone is activated to p-benzoquinone by the action of MPO (Smith et al., 1989; 

Rushmore et al., 1984; Low et al., 1995). However, the fact that administration of phenol does 

not produce the same effects as benzene is a problem in the phenolic metabolite hypothesis. A 

combination of phenol and hydroquinone is required to reproduce the hematotoxic effects of 

benzene (Eastmond et al., 1987). 

A ring-opening pathway of benzene metabolism leads to excretion of MA in the urine. 

The production of the highly reactive ring-opened dialdehyde MUC as an intermediate in the 

pathway to MA has been demonstrated in isolated hepatic microsomes; however, circulation of 

MUC in the blood has not been demonstrated (Witz et al., 1996). Administration of MUC 

reproduces the hematotoxic effects in mice; however, MUC is so reactive that it is not likely to 

reach the target tissues. The less reactive ring-opened metabolite CHO-M-OH has been detected 

in rat liver perfusate (Grotz et al., 1994) and has been shown to be hematotoxic in mice (Zhang 

26 
et al., 1995b). Thus, CHO-M-OH is another potentially toxic benzene metabolite, but there has 

been little research to define whether CHO-M-OH is quantitatively important. 

Benzene oxide was believed to be too reactive to escape the liver, but it has recently been 

measured in the bloodstream of mice, and its dose to the body may exceed that of pbenzoquinone (Lovern et al., 1997; Lindstrom et al., 1997). Thus, several metabolites of 

benzene may play a role in inducing the toxic effects of benzene, and a combination of several 

metabolites may be required to cause the full range of benzene-induced toxic responses. Large 

quantitative differences in the production of putative toxic benzene metabolites have been 

observed among different animal species (Sabourin et al., 1989, 1992). Because few data exist 

on the proportion of different benzene metabolites produced in humans, there is considerable 

uncertainty in selecting the most appropriate animal model for humans. 

There is evidence to support several different mechanisms by which benzene metabolites 

may cause hematotoxicity. Benzene metabolites form adducts with both proteins and DNA 

(McDonald et al., 1994; Lutz and Schlatter, 1977). Adduct formation with protein sulfhydryl 

groups inhibits tubulin polymerization during spindle formation and the activity of 

topoisomerase II in DNA replication and transcription (Irons and Neptun, 1980; Pfeifer and 

Irons, 1983; Frantz et al., 1996). These effects may explain the clastogenic activity of benzene. 

Benzene itself is not mutagenic; however, mutagenicity of several of its metabolites is well 

established (Dean, 1985). Benzene metabolites may also induce oxidative stress by depleting 

levels of GSH and by production of reactive oxygen species that react with cellular 

macromolecules (Snyder et al. 1987; Brunmark and Cadenas, 1988; Rao and Snyder, 1995). 

Benzene also disrupts production of the cytokine IL-1, which is essential for hematopoiesis. 

Niculescu et al. (1995, 1996) found that p-benzoquinone inhibits the activity of calpain, an 

enzyme that catalyzes the processing of pre-IL-1 to the mature cytokine. All of these effects 

may be involved in causing the toxic effects of benzene, but it is not possible to determine which 

of these effects is the primary mechanism for benzene toxicity. 

At low exposure levels, benzene is excreted primarily in the urine as sulfate or 

glucuronide conjugatges of phenolic metabolites or as MA (McMahon and Birnbaum, 1991; 

Sabourin et al., 1987). At oral doses of less than 15 mg/kg, 14C from benzene is nearly 

quantitatively recovered as metabolites in urine in rats and mice, but at a dose of 150 mg/kg, 

greater than 50% was cleared unmetabolized in the breath (Sabourin et al., 1987). Similarly, less 

than 6% of 14C from benzene was exhaled in the breath from inhalation exposure to 10 ppm (32 

mg/m3) in rats or mice, but at concentrations greater than 850 ppm (2718 mg/m3) rats exhaled 

48% and mice exhaled 14%. Urinary phenol and MA concentrations are correlated with 

benzene exposure level and can be used to monitor benzene occupational exposure (Ghittori et 

al., 1993). However, the database for evaluating the relative importance of different excretion 

pathways in humans is limited. Thus, there is a limited database for comparing the importance 

of human excretion pathways at varying dose levels with the results of experimental animal 

studies. 

Several different PBPK models have been developed to mathematically describe the 

uptake, distribution, metabolism, and excretion of benzene (Sato, 1988; Medinsky et al., 1989a, 

b, c; Travis et al., 1990a, b; Bois and Paxman, 1992). Each of these PBPK models has 

advantages, and some successfully simulate uptake, metabolism, and excretion in mice, rats, and 

humans. However, the utility of PBPK modeling to predict the dose of toxic metabolites to 

27 
target organs in humans is limited by the incomplete knowledge of the toxic metabolites and 

difficulties in identifying a suitable experimental animal model for humans. 

4. HAZARD IDENTIFICATION 

4.1. STUDIES IN HUMANS 

4.1.1. Oral Exposure 

Individual case reports of death from acute oral exposure to benzene have appeared in the 

literature since the early 1900’s. The benzene concentrations ingested by the victims often were 

not known. However, lethal oral doses for humans have been estimated at approximately 125 

mg/kg (Thienes and Haley, 1972). Accidental ingestion and attempted suicide with lethal oral 

doses of benzene have produced the following signs and symptoms: staggering gait, vomiting, 

shallow and rapid pulse, somnolence, and loss of consciousness followed by delirium, 

pneumonitis, collapse, and then central nervous system depression, coma, and death (Thienes 

and Haley, 1972). Lethality in humans has been attributed to respiratory arrest, central nervous 

system depression, or cardiac collapse (Greenburg, 1926). Ingestion of lethal doses may also 

result in visual disturbances or feelings of excitement and euphoria, which may suddenly change 

to weariness, fatigue, sleepiness, convulsion, coma, and death (Von Oettingen, 1940). 

4.1.2. Inhalation Exposure 

4.1.2.1. Hematotoxicity 

The recognition of benzene as etiologically significant in the development of aplastic 

anemia, a potentially life-threatening suppression of bone marrow activities, and pancytopenia, a 

reduction of the cellular elements of the peripheral blood, dates back to the 19th century 

(Goldstein, 1988). Exposure to benzene has also been associated with acute myelogenous 

leukemia (AML) and lymphoid malignancies in humans (Goldstein, 1988; Snyder, 1987), 

although the precise relationship between frank pancytopenia and the benzene-induced 

malignancies remains unclear (Snyder, 1987). That the hematopoietic abnormalities observed in 

humans exposed to benzene in an occupational setting have been reproduced in numerous animal 

experiments points to their universality. The comparatively low doses associated with their 

onset suggest that they may represent a sensitive effect of benzene toxicity. 

Human exposure to benzene occurs primarily via inhalation in the workplace, from 

gasoline vapors, tobacco smoke, and automotive emissions. Individuals exposed to benzene 

exhibit bone marrow depression, as evidenced by anemia (decreased RBC count), leukopenia 

(decreased WBC count), and/or thrombocytopenia (decreased platelet count). A depression of 

all three elements is called pancytopenia, and the simultaneous depression of RBCs, WBCs, and 

platelets, accompanied by necrosis of the bone marrow, is diagnostic of aplastic anemia. 

Patients with aplastic anemia also have exhibited mild bilirubinemia, changes in osmotic 

fragility of erythrocytes, shortened erythrocyte survival time, increased fecal urobilinogen, and 

mild reticulocytosis (Aksoy, 1991). The bone marrow profile may vary from aplasia to 

hyperplasia, and these symptoms may vary in frequency and severity in different patients. Other 

hematologic changes observed in humans chronically exposed to benzene include decreased 

28 
osmotic resistance in leukocytes, decreased phagocytic function of neutrophils, reduced 

glycogen content, and decreased activity of peroxidase of neutrophils and increased deltaaminolevulinic acid activity in erythrocytes and increased coproporphyrins in the urine (Aksoy, 

1991). 

As summarized in Table 1, the large number of epidemiologic studies that have explored 

the health consequences of benzene exposure in the workplace provide a considerable body of 

inferential evidence for the compound-related onset of hematopoietic effects in chronically 

exposed subjects. It should be recognized that for most of the studies there are uncertainties 

and/or variabilities in precisely delineating the concentrations of benzene to which subjects were 

exposed. For example, uncertainty surrounding the past or present workplace concentrations of 

the chemical may not permit the dose to which subjects were exposed to be estimated accurately. 

Even when an exposure term can be defined within an adequately cohesive range, its overall 

contribution to the understanding of the compound’s dose-response characteristics will probably 

be limited to a single data point from any given study. Thus, the utility of such a study might 

only arise through its consideration with others as a group, for example, in a “meta-analysis.” In 

such a survey, the establishment of a combined range for estimated exposures and their 

toxicological consequences can permit (1) the emergence of a unifying concept of the 

compound’s toxicity to humans and (2) the delineation of a range of concentrations or doses 

within which the compound’s toxicological threshold may reside. 

The paragraphs that follow describe the many epidemiologic studies that have helped to 

establish the onset of benzene’s hematologic effects. 

In general, it is assumed that for the occupational exposure studies under consideration, 

the levels of exposure to benzene were greater in earlier studies than in more recent studies, a 

reflection of times when occupational exposure standards and guidelines were less rigorous. For 

example, studies that have examined occupational exposure to benzene at relatively high 

concentrations include several that evaluated hematologic surveillance records of 459 rubber 

hydrochloride workers from the Pliofilm production departments of a rubber products 

29
Table 1. Hematotoxicity of benzene—occupational exposure 

Population/Industry 

No. of 

Subjects Exposure Effects Reference 

Rotogravure printers 6 77–3386 mg/m3 benzene Pancytopenia Erf and Rhoads, 1939 

Rotogravure printers, New York 

332 35–3386 mg/m3 benzene 23 cases of significant cytopenias Goldwater, 1941; Goldwater 

and Tewksbury, 1941; 

Greenburg et al., 1939 

Shoe factory, Finland 147 1278 mg/m3 benzene for ~10 yr 107 cases of hematologic abnormalities Savilahti, 1956 

Shoe factory, Finland 147 Followup of Savilahti (1956) study, 9 

years later 

Persistent cytopenias; one death from acute leukemia 

Hernberg et al., 1966 

Rubber factory, Ohio 1104 1597 mg/m3 (mean, ~319 mg/m3) 

benzene 

83 cases of mild hematologic abnormalities, 25 

cases of more severe pancytopenia (9 required hospitalization, 3 of whom died) 

Wilson, 1942 

Watch industry 216 No exposure data; workers followed for 

10 years after cessation of exposure to benzene 

Four cases of persistently decreased blood counts; 

one death from aplastic anemia 9 years after 

cessation of exposure 

Guberan and Kocher, 1971 

Rubber-coating industry 

< 80–399 mg/m3 benzene (prior to 

installation of control measures) 

Follow-up suggested mild persistent anemia Pagnotto et al., 1961 

Shoe manufacturing, Turkey Controls (healthy hospital workers) 

217 100 

96–671 mg/m3 benzene for 3 months–17 

years 

51 cases of hematologic abnormalities, including 6 

cases of pancytopenia 

Aksoy et al., 1971 

Various industries using benzenecontaining adhesives, Turkey 

32 diagnosed with blood dyscrasias 

479–1597 mg/m3 benzene for 4 

months–15 years 

32 cases of significant aplastic anemia; eight deaths from thrombocytopenic hemorrhage and infection 

Aksoy et al., 1972 

Various industries using benzenecontaining adhesives, Turkey 

44 diagnosed with blood dyscrasias 

Follow-up study over 2–17 years after last 

exposure to 479–2077 mg/m3 benzene for 

4 months–15 years 

Complete remission in 23 (52%), death due to complications of pancytopenia in 14 (32%), 

leukemia in 6 (14%), myeloid metaplasia in 1 (2%) 

Aksoy and Erdem, 1978 

30
Table 1. Hematotoxicity of benzene—occupational exposure (continued) 

Population/Industry 

No. of 

Subjects Exposure Effects Reference 

Turkish, various 231 Exposure to solvents, thinners, and 

similar materials in 40 small industries; in one facility, benzene concentrations 

averaged 351 mg/m3; in the remainder, 

benzene levels < 1 ppm 

351 mg/m3: two cases of acute myeloblastic 

leukemia in facility where levels measured 351 mg/m3 

< 1 ppm: ild abnormalities in 14 workers (6.1%); 

differential diagnosis of 186 workers, monocytosis in 8, eosinophilia in 5, and basophilia in 3; one case each of acute myeloblastic leukemia, Hodgkin’s 

disease, and lymphoma 

Aksoy et al., 1987 

Chemical production facility 

282 men Jobs assigned to exposure categories: 

(1) < 6 mg/m3 TWA 

(2) 6–29 mg/m3 TWA (3) 32–77 mg/m3 TWA (4) 80– ~96 ppm TWA exposures ranged from < 1 years to > 20 years 

Slight statistically significant (p<0.05) (but not biologically significant) decreases in bilirubin and RBC counts; no correlations between peripheral blood counts and latency, duration, or intensity of 

benzene exposure 

Townsend et al., 1978 

Refinery 303 men Exposure to benzene levels of 1.70 mg/m3 

(mean) for an average of 8 years for white 

employees and 4.5 years for nonwhite 

employees); average length of followup, 13 years 

No significant changes in hematology of workers as a group; one case of multiple myeloma; one death 

from multiple myeloma and one from malignant 

melanoma 

Tsai et al., 1983 

Chinese shoemaking and printing 

300 Mean TWA exposures: 

Benzene—105 and 188 mg/m3 (55 and 65 

months) for men and women, respectively; Toluene—173 and 154 mg/m3 (95 and 73 

months); Mixture—45 mg/m3 benzene + 68 mg/m3 

toluene and 45 mg/m3 benzene + 79 

mg/m3 toluene (159 and 120 months) 

Changes in absolute and/or relative peripheral cell counts (statistically significant at p<0.01 or p<0.05): Benzene—decreased lymphocyte counts (women; 

men and women, combined); eosinophilia (women; 

men and women combined); decreased monocytes (women) Toluene—decreased lymphocyte counts (women; 

men and women, combined); eosinophilia (all 

groups) Mixture—decreased lymphocyte counts (men; 

women; men and women combined); significant eosinophilia (men and women combined) 

Yin et al., 1987b 

m

31
Table 1. Hematotoxicity of benzene—occupational exposure (continued) 

Population/Industry 

No. of 

Subjects Exposure Effects Reference 

Rubber industry 

Workers Unexposed controls Workers in various Chinese industries Low-exposure group Below median Above median Unexposed controls Rubber industry 

459 Mean estimated exposure from 1948 to 

1975, 48–64 mg/m3 

Significant increases in WBCs, RBCs, and Hgb during the 1940’s (Pearson correlations: r = 0.50 for WBCs, r = 0.44 for RBCs, and r = 0.71 for Hgb), but not during the ensuing 25 years; rapid decline in exposure for the workers consistent with increases in blood count values 

Kipen et al., 1988, 1989 

200 268 

0.03–4.5 mg/m3 8-hour TWA over a 10-

year period 

No abnormal hematology Collins et al., 1991 

11 22 22 44 

Workers exposed to benzene, but not to other solvents, at TWA of 24 (low exposure), 44 (below median), or 294 (above median) mg/m3 for an average of 

6.3 years; doses were measured by personal dosimetry for 2 weeks 

Significant increase in mean corpuscular volume 

and significant decreases in ALC, WBCs, RBCs, HCT, and platelets in above-median group; significant reductions in ALC, RBCs, and platelet count in below-median group; but only reduction in ALC was significant in low-exposure group 

Rothman et al., 1996a 

657 Estimates of benzene exposure (109 

mg/m3) for last 30, 90, or 180 days before 

blood tests were correlated with hematologic data 

Weak correlation between RBC and benzene exposure, but strong correlation between low WBC counts and benzene exposure, with no evidence of a threshold 

Ward et al., 1996 

32
manufacturer in Ohio (e.g., Kipen et al., 1988, 1989; Ward et al., 1996). The data indicate that 

during the time period 1940–1948, significant increases in erythrocyte, Hgb, and leukocyte 

levels were observed among the workers, with the responses displaying a positive time-related 

trend (1940–1948). Thus, overall cell numbers and values appear to have been higher at the later 

time points, the prevailing levels of exposure to benzene becoming reduced as industrial hygiene 

practices improved. Kipen et al. (1988, 1989) compared the utility of two exposure assessment 

models for correlating the observed hematologic consequences with plausible estimates of 

benzene exposure in the absence of definitive monitoring data at the earlier time points. Thus, 

between 1948 and 1975, when the workers were exposed to mean 8-hour TWA concentrations of 

only 15–20 ppm (48–64 mg/m3

), no significant relationships between benzene levels and any 

hematologic parameters were observed. By contrast, significantly lower cell counts and 

hemoglobin concentrations were observed in exposed subjects, where the levels of benzene 

exposure were thought to be as high as 75 ppm (240 mg/m3) at the earlier time points (pre-1948) 

(Kipen et al., 1989). 

Cody et al. (1993) observed significant hematologic effects, including reduced RBC and 

WBC counts in 161 male rubber workers exposed to median peak concentrations (i.e., only the 

peak concentrations for any given exposure time were reported) of 30–54 ppm or more for a 12-

month period during 1948. The 30 ppm value was considered a 1-year LOAEL for hematologic 

effects. In this rubber plant, workers who had blood dyscrasias were excluded from working in 

the high-benzene units. Furthermore, individual workers who had more than a 25% decrease in 

WBC counts from their pre-employment background count were removed from the high-benzene 

units and placed in other units with lower benzene concentrations. Sensitive individuals 

therefore could have been excluded from the analysis. The 30 ppm value is the low end of the 

range of median values (30–54 ppm) reported by Crump and Allen (1984) and used in the Kipen 

et al. (1988) and Cody et al. (1993) studies. These are the same Pliofilm workers studied by 

Rinsky et al (1981, 1987). 

The analysis of Ward et al. (1996) differs from the earlier investigations of the same 

rubber worker cohort by Kipen et al. (1988, 1989) and Cody et al. (1993) in that a nested casecontrol design was used. Incident cases were defined as the first occurrence of a low WBC or 

RBC count, and matched controls were chosen from those tested within 6 months of the case’s 

blood test date. In contrast to the earlier analyses of hematologic screening data from the same 

plant, the Ward et al. (1996) study used the entire data set, evaluated the exposure-response 

relationship on the basis of individual dose metrics, and controlled for the temporal trends in preemployment blood count screening. 

The Ward et al. (1996) analysis used hematologic screening data for 657 of 1037 

individuals employed at the plant from 1939 through 1976. There was a total of 21,710 blood 

test records. The study used a case-control design and estimated benzene exposures using the 

job exposure matrix developed by Rinsky et al. (1981). The effects of benzene exposure in the 

30, 90, and 180 days before the blood test date, as well as cumulative exposure until the blood 

test date, were examined using conditional logistic regression. For WBC count there was a 

strong exposure response, and all of the exposure metrics selected showed a significant 

relationship with low blood count. For RBC count there was a weak positive exposure response 

that was 

33 
significant (p<0.03) for one of the dose metrics. The maximum daily benzene exposure estimate 

in this study was 34 ppm (109 mg/m3). There was no evidence for a threshold for the 

hematologic effects of benzene exposure, suggesting that even exposure to relatively low levels 

(e.g., < 5 ppm [16 mg/m3

]) could result in hematologic suppression. The results of this study are 

consistent with test animal studies that have demonstrated a decrease in peripheral lymphocyte 

counts at benzene exposures as low as 10 ppm (32 mg/m3) and a stronger effect of benzene 

exposure on lymphocytes than on red cells; therefore, if one assumes that Ward’s Pliofilm 

workers exposure estimates are correct, they suggest that hematologic effects of benzene occur 

even at 5 ppm. Because of controversy surrounding the exposure estimates used by Ward et al., 

this finding could be biased. 

The nature of the problem is that the estimates of early exposure to benzene (prior to 

1946) are not based on any measured ambient air inhalation data for any of the more than 1200 

Pliofilm workers who begn work in the plants prior to 1946. Estimates of exposure to benzene 

in the set of exposure indices chosen by Ward and her associates (1996) were based on the 

exposure scenario described by Rinsky et al. (1981), which basically assumed that the exposures 

were the same as the local and State standards established for benzene at the time (in the years 

prior to 1946). This assumption has been challenged by others (Crump, 1992; Paustenbach et 

al., 1992), who claim that the exposures during this time were much higher. Thus, the 

assumption that the exposures could have been as low as < 5 ppm may not be correct. The 

weaknesses in the Ward et al. study are discussed further in U.S. EPA (1998a). No LOAEL or 

NOAEL could be established for this study. 

In another occupational study that examined the health effects of benzene, Fishbeck et al. 

(1978) studied the hematologic characteristics of 10 Dow Chemical employees who had been 

exposed to high concentrations of the chemical in an ethyl cellulose sheeting operation for 

various durations between 1937 and 1965. Duration of exposure was from fewer than 4 to nearly 

30 years, with average 8-hour TWA benzene concentrations exceeding 25 ppm (80 mg/m3

) for at 

least a part of the period of employment. As described by Fishbeck et al. (1978), in a 1953 

industrial hygiene survey, workers who had an “operator” job classification had the highest 

exposure to the compound, with an 8-hour TWA of 30–35 ppm (96–112 mg/m3

) and occasional 

“spikes” of up to 937 ppm (2990 mg/m3

). However, a 1963 study described by Fishbeck et al. 

(1992) revealed 8-hour TWAs ranging from 37 to 132 ppm (118 to 422 mg/m3), with four 

employees’ exposure levels exceeding 100 ppm TWA (319 mg/m3

). Concurrent hematologic 

evaluations undergone by these employees included Hgb, HCT, RBC and WBC counts, 

sedimentation rates, platelet counts, differential blood counts, clot retention determinations, and 

blood indices. 

Key findings included enlarged RBCs, transient anemia, reduced hemoglobin 

concentrations, and mean corpuscular volumes (MCVs) exceeding medically accepted values. 

However, the authors noted that since 1963, the employees’ MCVs had shown a reduction from 

the highest values obtained, such that 5 of the 10 employees had MCVs within the normal range. 

From these data, the authors concluded that at least some of benzene’s effects on the 

hematopoietic system may be reversible. This conclusion may be premature. This small group 

of 10 employees was a survivor population and therefore may have introduced a potential bias. 

They were selected because they were listed on current employment rolls of the company when 

the study was in progress. Employees who had left employment earlier, perhaps with a benzene34 
related nonreversible condition, could not be selected for inclusion in this study. Therefore, 

persons with an irreversible condition due to exposure to benzene could not have been 

represented in this group of 10. This study cannot be used to establish a NOAEL or a LOAEL. 

Aksoy et al. (1971) assessed hematologic abnormalities in 217 apparently healthy men 

exposed to benzene in the shoe manufacturing industry. The test battery included RBC, WBC, 

platelet, and differential cell count; packed cell volume (PCV); bone marrow examination; and 

Hgb analysis. The subjects, 12–58 years old, were compared with 100 unexposed, age-matched 

hospital workers. Benzene levels ranged from 15 ppm (48 mg/m3

) during nonworking hours to 

210 ppm (671 mg/m3

) during the use of adhesives containing benzene; exposure duration ranged 

from 3 months to 17 years. A total of 217 exposed workers were involved in the study, with 51 

(23.5%) displaying hematologic abnormalities. Of these, leukopenia was present in 9.7%, 

thrombocytopenia in 1.8%, leukopenia associated with thrombocytopenia in 4.6%, pancytopenia 

in 2.8%, acquired pseudo-Pelger Huet anomaly in 0.5%, eosinophilia in 2.4%, and basophilia in 

0.5%. 

Bone marrow examination of 11 of the 51 workers with abnormal peripheral blood cell 

counts revealed (1) normocellularity with no abnormal histology (two patients), (2) slight 

hypocellularity (three patients), (3) hypercellularity (one patient), or (4) normocellularity with 

abnormal histology. Abnormal findings present in various combinations included maturation 

arrest in the erythroid and myeloid series, maturation arrest in the granulocytic series, and 

marked vacuolization in the myeloid series. One worker with normal hematology developed 

acute erythroleukemia 4 years later. The investigators concluded, based on the incidences of the 

various blood abnormalities, that benzene had a greater effect on leukocytes (with basophilia and 

eosinophilia as inconsistent findings) than on platelets. In a follow-up study, Aksoy et al. (1972) 

reported that 8 of 32 workers diagnosed with pancytopenia died, mainly from infection and 

bleeding. 

In an industrial hygiene survey of 282 Dow Chemical workers who were categorized by 

job classification into four groups on the basis of their exposure to benzene, exposed individuals 

were monitored for a number of clinical chemistry and hematologic parameters and compared 

with an age- and sex-matched cohort of unexposed Dow employees serving as controls 

(Townsend et al., 1978). Ranges of benzene concentrations within the groups were < 2, 2–9, 

10–24, and > 24 ppm (< 6.4, 6.4–29, 32–77, and > 77 mg/m3

). However, as noted by the 

authors, neither the RBC counts nor any other hematologic parameter correlated with exposure 

intensity, duration of exposure, or estimated career dosage (Townsend et al., 1978). In fact, no 

consistent relationship between any monitored parameter and benzene exposure was apparent 

from their data. 

A report by Yardley-Jones et al. (1988, 1990) described a survey in which blood samples 

were obtained from 66 refinery workers exposed to benzene at comparative low concentrations 

and 33 unexposed employees serving as controls. Exposure levels of < 10 ppm (32 mg/m3

) were 

estimated through consideration of the exposed employee’s workstation, supplemented by a 

small amount of personal monitoring data. Exposed subjects were assessed for the appearance of 

sister chromatid exchange in peripheral lymphocytes and the ability of a subject’s 

-glucuronidase/sulfatase-treated urine to promote gene reversion in the Ames test. In addition, 

some hematologic parameters were measured in whole blood, although the largely negative 

35 
findings gave no indication of a cause-effect relationship for benzene within the estimated 

exposure range. 

Bogadi-Sare et al. (2000), in a medical surveilance study, examined 49 female workers in 

the shoemaking industry who were exposed to solvent mixtures and 27 nonexposed controls. 

Exposures were as high as 15 ppm benzene and 50 ppm toluene. Significant differences in the 

levels of benzene and toluene in blood and phenols in postshift urine between the exposed and 

the control group confirmed solvent exposure. The number of B-lymphocytes was lower in the 

shoe workers than in the controls (p=0.01). Significant correlations were found between the 

level of immunoglobulin G and benzene in the work atmosphere. Confounding factors appeared 

to play no role in the immunological findings. The authors found that exposure to benzene 

concentrations lower than 15 ppm can induce depression of the circulating B-lymphocytes. 

Khuder et al. (1999) reported on a group of 105 petroleum workers exposed to 0.14–2.08 

ppm (0.45–6.6 mg/m3

) benzene who, over time, had small but statistically significant falls in 

certain blood counts. However, there were problems with this study, including a decrease in the 

red cell MCV, a finding contrary to what is observed in benzene toxicity (Goldstein and Cody, 

2000). 

Collins et al. (1991) used a cross-sectional study design to compare five hematologic 

parameters in 200 benzene-exposed workers with those of 268 unexposed employees in the same 

plant. The type of work is not identified in the study. Estimates of the 8-hour TWA benzene 

concentration were obtained from monitoring information coupled with knowledge of the 

applicable work assignment and professional judgment. Benzene concentrations ranged from 

0.01 to 1.4 ppm (0.03–4.5 mg/m3

) for exposed subjects. Parameters under investigation included 

RBCs, WBCs, Hgb concentration, platelets, and MCV. The authors noted statistically 

significant differences in some of the hematologic parameters, based on demographic and 

lifestyle factors. Thus, when multiple regression analyses were applied using the confounding 

factors and current exposure as independent variables, no significant correlations between 

benzene exposure and any hematologic parameter was evident. Again, this was a study of 

currently employed workers and, as such, represented a healthy group of individuals, who, 

having survived potentially adverse effects from exposure to benzene and other environmental 

hazards, were able to continue to work and be included in this cross-sectional study. Hence, this 

study cannot be used to establish a LOAEL. 

Tsai et al. (1983) examined mortality patterns in a cohort of 454 former and current male 

employees who were selected by industrial hygienists from a larger group of 20,000 present and 

former refinery employees. The cohort members were chosen because they had worked directly 

in the benzene, ethylene, aromatic distillate hydrogenation, or caiman units, the principal 

petrochemical units in the refinery, for 1 to 21 years beginning in 1952. Average length of 

“exposure” was 7.4 years. Maintenance workers who were assigned on a regular basis to these 

same units were also included. Irregularly assigned workers to these units were not included, 

although they may have been exposed to benzene. Personal exposures to benzene were 

determined from 1979 to 1982. The median air concentration was 0.53 ppm in the “benzenerelated units,” while in the refinery as a whole, the median was 0.14 ppm. In nonbenzene 

exposure areas, the median was 0.07 ppm. This one-time “snapshot” of benzene exposure levels 

says little about what the exposures may have been in the past or about study duration, nor do 

36 
they tell us whether these personal samples relate to the 303 medically surveyed workers 

discussed below, as they are a different group from the 454 who were part of the mortality 

cohort. From this study it is impossible to determine the concentrations of benzene to which the 

medically surveyed workers were exposed. 

The analysis of overall mortality in the mortality cohort of 454 revealed no excesses. In 

fact, mortality from all causes and from diseases of the circulatory system was significantly 

below expected values, based on comparable groups of U.S. males. The authors concluded that 

the presence of a healthy-worker effect produced significantly lower risk estimates, all causes 

(standard mortality ratio = 0.53, p<0.05); only 34 deaths occurred, whereas 58.7 were expected. 

The authors noted the relative youthfulness of the cohort. An internal but separate analysis of a 

comparison group of 823 people consisting of 10% of the workers who were employed in the 

same plant in operations not related to benzene produced higher relative risks of 0.90 and 1.31 

for all causes and cancer at all sites, respectively (p<0.28 and p<0.23), higher than those of the 

so-called “benzene-exposed” workers. Because of the appearance of significantly reduced risks 

in this preselected cohort, the possibility of bias in the selection process should not be ruled out. 

In addition, hematologic parameters were analyzed in a separate group of 303 male 

“benzene workers” (according to the supervisors who selected them) beginning in 1959 and 

ending in 1979. During this time, 11 deaths occurred in this group, whereas 21.5 were expected. 

Up to four hematologic tests per year were conducted on just these workers. Total and 

differential WBC counts, Hgb, HCT, RBC counts, platelets, and clotting times were calculated, 

and a 95% range and standard deviation were derived for each hematologic test. Unfortunately, 

as the authors pointed out, the data could not be compared with any “healthy population” 

standard, because existing “normal” ranges are derived largely from hospital experiences. The 

authors suggested that, based on the experience of this program, there is a need to establish 

different interpretation criteria for screening for adverse effects of benzene exposure. 

Thus, the Tsai et al. (1983) study is similar to the studies by Collins et al. (1991), 

Fishbeck et al. (1978), and Yardley-Jones et al. (1988, 1990) in that the study population was a 

currently employed group of workers (except for the 11 who died after the medical surveillance 

was begun) and, as such, a survivor population in which few or no adverse health effects would 

have been found. In addition, these medically surveyed employees were healthy and young. 

They had been exposed for a short time to minimal levels of benzene that generally did not 

exceed 0.53 ppm and probably were a lot less, based on survey results. Therefore, the selection 

of a NOAEL is not considered on the basis of this study. 

Considering data from the studies that have addressed the possible toxicological 

consequences of exposure to benzene, a continuum of hematologic responses that are 

etiologically linked to exposure to benzene is evident, although the precise threshold level for 

these effects remains uncertain. However, some clarification of this issue has been provided by 

an ongoing collaboration between the National Cancer Institute and the Chinese Academy of 

Preventive Medicine that has generated a series of reports on both large and small cohorts of 

subjects who were occupationally exposed to benzene. 

Data from China describe clinical aplastic anemia in factories with exposure levels said 

to range from 93 to 1156 mg/m3

. Yin et al. (1987b) found 2676 cases of benzene 

37 
poisoning—defined as a WBC count less than 4000 cells /mm3

 and a history of benzene 

exposure—from a review of more than 500,000 benzene-exposed workers in China. The 

geometric mean concentration in 50,255 workplaces was 18.1 mg/m3

. It was reported that 64.6% 

of the workplaces had benzene exposure levels less than 40 mg/m3

. From their review, the 

authors concluded that cases of benzene poisoning can occur even in factories with less than 40 

mg/m3 concentration of benzene. 

Dosemeci et al. (1996) discussed the findings of a large-scale epidemiologic study 

involving subjects employed between 1949 and 1987 in 672 factories in 12 cities in China. The 

authors sought to establish plausible estimates of the levels of benzene exposure through an 

industrial hygiene-based survey of all factory/work unit/job title/calender period data, providing 

assessments that were subsequently correlated with one of six hypothetical exposure ranges 

between < 1 and > 50 ppm. Thirty-eight percent of a total of 18,435 exposure estimates were 

confirmed by workstation monitoring. The authors set hematologic criteria that defined benzene 

poisoning to include a WBC count of < 4000 cells/mm3

, platelets < 80,000/mm3, etc., thereby 

allowing the identification of a case-defined subset of individuals who could be matched with 

estimated levels of benzene exposure as described above. Relative risks of benzene poisoning 

by intensity of exposure during the most recent 1.5-year exposure period compared with those 

exposed to < 5 ppm (16 mg/m3

) yielded estimated risks of 2.2 (95% CI = 1.7–2.9), 4.7 (95% CI = 

3.4–6.5), and 7.2 (95% CI = 5.3–9.8) for exposure levels of 5–19 ppm (16–61 mg/m3

), 20–39 

ppm (64–125 mg/m3

), and > 40 ppm (> 128 mg/m3), respectively. Clearly, if a significantly 

elevated risk of benzene poisoning is an indication of hematotoxicity, then certainly exposures to 

benzene at 5–19 ppm are hematotoxic. The decrease in ALC contributed substantially to the 

decrease in WBC count, based on the results of the Rothman et al. (1996a) study discussed in the 

next review. In general, recent exposure history appeared to be more closely related to elevated 

risk than exposure duration, suggesting the comparatively rapid onset of benzene hematotoxicity. 

These results support a threshold of benzene hematotoxicity in humans in the 5–19 ppm range, in 

broad agreement with the emerging exposure-response range that is apparent from the 

epidemiologic studies described in this section. 

Some data that have shed further light on this issue have been provided by the same 

group (Rothman et al., 1996a) that conducted a small cross-sectional study that compared 44 

workers exposed to a range of benzene concentrations similar to those of Dosemeci et al. (1996) 

with 44 age- and gender-matched unexposed controls. The purpose of this study was to show 

that exposure to benzene affects all the major peripheral blood elements, with the ALC being the 

most sensitive indicator of benzene-induced hematotoxicity. From public health district records, 

workers from three workplaces in Shanghai, China, where benzene was used as a solvent and 

unexposed workers from two workplaces in the same geographic area that did not use benzene 

were selected as test subjects and controls, respectively. The three workplaces using benzene 

included a factory that manufactured rubber padding for printing presses, a factory that 

manufactured adhesive tape, and a factory that used benzene-based paint. The control 

workplaces included a factory that manufactured sewing machines and an administrative facility. 

Subjects who had a history of cancer, therapeutic radiation, chemotherapy, or current pregnancy 

were excluded. Requirements for inclusion in the study were current employment for at least 6 

months in a factory that used benzene, minimal exposure to other aromatic solvents, and no 

exposure to other known bone marrow-toxic chemicals or ionizing radiation. Controls who had 

38 
no history of occupational exposure to benzene or other bone marrow-toxic agents were 

frequency-matched to the exposed subjects on age (5-year intervals) and gender. 

Benzene exposure was monitored by organic vapor passive dosimetry badges worn by 

each worker for a full workshift on 5 days within a 1- to 2-week period prior to collection of 

blood samples. Benzene exposures of controls in the sewing machine factory were monitored 

for 1 day, but no exposure monitoring was performed in the administrative facility. Compound 

exposure was also evaluated by analyzing for benzene metabolites in urine samples collected at 

the end of the benzene-exposure period for the exposed subjects. Median levels of the urine 

metabolites phenol, catechol, MA, and hydroquinone correlated positively with the 8-hour TWA 

personal air level measurements of benzene by exposure category (Table 2). 

It is possible that air exposure to benzene may have been overestimated among workers 

wearing marginally effective respirators. However, a comparison of urinary phenol levels to the 

air benzene levels can be made. Urinary phenol is thought to increase in a linear fashion up to 

100 ppm benzene. Roughly, 50 μg/mg creatinine is thought to be equivalent to an 8-hour TWA 

of 10 ppm benzene. For 19 workers exposed to more than 31 ppm who had an air and urine 

sample taken on the same day, the median benzene exposure was 73 ppm and the median urine 

phenol level was 351 μg/mg creatinine, about what one would predict (letter from Nathaniel 

Rothman, National Institutes of Health, Bethesda, MD, to David Bayliss, NCEA-W, July 28, 

1999). This is a test of the validity of the benzene air sampling effort. Dosemeci et al. (1996) 

estimated that only about 5% of more than 1000 workers who were exposed to 100 ppm benzene 

or higher in their factories were diagnosed with benzene poisoning. It is not known whether 

severe hematotoxicity occurred to that 1 individual in 19 who was reported to have a TWA of 

328.5 ppm. It is not always possible to obtain a perfect correlation between a surrogate measure 

of exposure and an effect. 

Historical benzene exposure of the subjects was evaluated by examining employment 

histories. However, this information was not used to classify subjects by category of current 

exposure. Data on age, gender, current and lifelong tobacco use, alcohol consumption, medical 

history, and occupational history were collected by interview. 

Six hematologic parameters were evaluated: total WBC count, ALC, HCT, RBC count, 

platelet count, and MCV. Total WBC counts and ALCs were performed using a Coulter T540 

blood counter. Abnormal counts were confirmed by hand. The Coulter company estimated that 

the machine used in this study counts about 10,000 WBCs to generate a total WBC count and the 

percent lymphocytes, which are multiplied to provide the ALC. The hand count involves 

counting a standard 100 cells. The manual count also used to generate the percent lymphocytes 

in this study was the standard 100 cells. Much higher-quality data would be expected for the 

former and less precision for the latter, due to sampling variation (as well as human scoring 

error). However, the Coulter counting machine actually generated larger and more significant 

39
Table 2. Median benzene urine metabolites, by exposure category, in a 

study of workers exposed to benzene in Shanghai, China, 1992a 

Exposure Category Phenol Catechol MA Hydroquinone 

Control (N = 44) 17.3 3.2 0.18 1.6 

Exposed (< 31 ppm, N = 22) 38.9 7.0 8.15 12.8 

Exposed (> 31 ppm, N = 22) 348.7 65.7 46.8 64.3 

a

Units μg/mg creatinine for all metabolites. 

Source: Adapted from Rothman et al., 1996a. 

differences in the ALC between the exposure groups than did manual counts. The authors 

reported that previous studies that reported ALC and percent lymphocytes based on manual 

counts may have underestimated group differences due to nondifferential measurement error. 

Benzene metabolites in urine were measured by an isotope dilution gas chromatography/mass 

spectrometry assay. Correlation analyses were completed using the Spearman rank order 

correlation method. The Wilcoxon rank sum test was used to test for hematologic differences. 

Twenty-one of the 44 subjects in both the exposed and control groups were female. 

Mean standard deviation years of occupational exposure to benzene was 6.3 (standard 

deviation = 4.4 years) with a range of 0.7–16 years. The median 8-hour TWA benzene exposure 

concentration for the 44 exposed workers was 31 ppm (99 mg/m3

). The 8-hour TWA for each 

subject was derived from the geometric mean of the five air measurements that were gathered on 

each of the exposed workers on the five different days during which the samples were taken. 

Exposure to toluene and xylene was less than 0.2 ppm (0.6 mg/m3) in all exposure categories. 

The exposed group was then subdivided into two equal groups of 22 workers consisting of those 

exposed to greater than the median concentration of 31 ppm versus those exposed to less than the 

median concentration according to the 8-hour TWA. The median (range) 8-hour TWA exposure 

concentration was 13.6 (1.6–30.6) ppm (43.4 [5.1–97.8] mg/m3

) for the low-exposure group and 

91.9 (31.5–328.5) ppm (294 [101–1049] mg/m3

) for the high-exposure group. A subgroup of the 

low-exposure group composed of 11 individuals, none of whose five separate measurements 

exceeded 31 ppm (100 mg/m3

) at any time during the monitoring period, was also analyzed for 

ALC. The median (range) 8-hour TWA exposure of these individuals was 7.6 (1–20) ppm (24 

[3.2–64] mg/m3

). All five air measurements were completed within 1 to 2 weeks of the 

phlebotomies on all subjects. 

All six blood elements measured were significantly different in the highest benzene 

exposure group when compared with controls (Table 3). ALC, WBCs, RBCs, HCT, and 

platelets were all significantly decreased, and MCV was significantly increased with increasing 

benzene exposure. These effects are consistent with the hematotoxic effects of benzene shown 

in Aksoy (1989) and Goldstein (1988). ALC was reduced from 1.9 × 103

/μL blood in controls to 

40
Table 3. Comparison of mean peripheral blood counts (with standard 

deviations), by exposure status, in a study of workers exposed to benzene in 

Shanghai, China, 1992 

Exposure 

Category 

WBC Count 

(×103/L blood) 

ALC 

(×103/μL blood) 

RBC Count 

(×106/L blood) 

4.7 (0.6) 

4.6 (0.6)b 

4.2 (0.6)c 

HCT 

(%) 

MCV 

(m3) 

Platelets 

(×103/L blood) 

Control 

(median = 

0.02 ppm, 

N = 44) 

Exposed 

( 31 ppm, 

N = 22) 

Exposed 

(> 31 ppm, 

N = 22) 

a

p<0.01 b

p<0.05 c

p<0.001 

6.8 (1.7) 

6.4 (1.8) 

5.6 (1.9)a 

42.0 (5.6) 88.9 (4.9) 

41.2 (5.7) 89.8 (3.9) 

38.8 (5.3)b 92.9 (3.4)c 

166 (59) 

132 (45)b 

121 (43)a 

1.9 (0.4) 

1.6 (0.3)a 

1.3 (0.3)c 

Source: Adapted from Rothman et al., 1996a. 

1.6 × 103/μL (p<0.01) in the < 31 ppm (99 mg/m3) group and to 1.3 × 103/μL (p<0.001) in the 

group exposed to > 31 ppm (99 mg/m3

) benzene. A dose-response relationship is evident in five 

measures of hematotoxicity. The RBC and platelet counts were also significantly reduced in the 

< 31 ppm (99 mg/m3

) exposure group, but only ALC was significantly different in the lowestexposure subgroup (median = 7.6 ppm, 24 mg/m3

). In this subgroup of 11 workers, whose 

median 8-hour TWA exposure was 7.6 ppm (24 mg/m3

) benzene, ALC (1.6 × 103/μL) was 

similar to that of the larger group of 22 who were exposed to < 31 ppm benzene. Thus, ALC is 

the most sensitive measure of benzene hematotoxicity, representing a sentinel of a continuum of 

the likely hematotoxic consequences of prolonged or enhanced exposure to benzene below 31 

ppm. 

One feature of the Rothman et al. (1996b) study was its attempt to probe the mechanisms 

by which benzene may bring about the hematotoxicological symptoms that were evident in the 

blood samples of exposed subjects. For example, although most peripheral blood cell levels 

were decreased in exposed workers as compared with controls, peripheral cytokine levels (IL-3, 

IL-6, erythropoietin, granulocyte colony-stimulating factor, and tissue necrosis factor-alpha) 

were similar in exposed workers and controls, suggesting that benzene may not affect these 

growth factor levels in peripheral blood. By contrast, when the rates of mutations were 

measured at the glycophorin A (GPA) locus, an increased frequency of somatic cells with the 

NN variant was interpreted as an indication of the compound’s capacity to produce geneduplicating but not gene-inactivating mutations at the GPA locus in bone marrow cells (Rothman 

et al., 1995). 

41
The study by Rothman et al. (1996a) is one of few occupational studies in which 

individual- and workstation-monitored benzene concentrations provide a range of values that 

encompass the exposure-response area into which, as judged by the data from other larger-scale 

epidemiologic studies, the point-of-departure for the compound’s hematologic effects appears 

likely to fall. Further marking the study’s relevance as a basis for developing quantitative 

reference values for the compound is the exposure-response relationship that emerged between 

ALCs and benzene concentrations. This observation is based on the inverse correlations 

between these parameters that were reported for benzene-exposed workers and on the exposure 

response that was observed when the data for exposed subjects were grouped on either side of 

the overall median benzene concentration of 31 ppm (99 mg/m3

). 

Further assurance that the study met sufficient criteria for its use in developing an 

inhalation RfC for benzene is provided by the facts that worker exposure to toluene and other 

potentially harmful solvents was minimal in the exposed group and that the study protocol 

fulfilled adequate criteria for age- and sex-matching of exposed and control subjects. 

As discussed above, a number of subgroups of benzene-exposed workers have been 

defined in the Rothman et al. (1996a) study, including (1) a high-dose subset of 22 workers, the 

means of whose 5-day passive dosimetry measurements for benzene were greater than the 

median (31 ppm [99 mg/m3

]) of the 5-day mean benzene exposure concentration values for the 

benzene-exposed group as a whole; (2) a low-dose group whose 5-day mean TWA benzene 

exposure values were lower than 31 ppm (99 mg/m3

) (median benzene concentration = 13.6 ppm 

[43.4 mg/m3

]); and (3) a subset of the low-dose group whose exposure levels in all 5-day 

consecutive breathing zone measurements were lower than 31 ppm (99 mg/m3

) (median benzene 

concentration = 7.6 ppm [24.3 mg/m3

]). Because a reduction in ALC was the only significant 

hematological effect obtained for all of the benzene groups, reduction in ALC was designated as 

the critical effect, and the lowest exposure median of 7.6 ppm was defined as the LOAEL. There 

is some concern in selecting the 7.6 ppm (24.3 mg/m3) concentration as the point of departure for 

RfC development, because the subjects in question (N = 11) may provide an inadequate groupwise comparison to the entire control group (N = 44). Nonetheless, a significant reduction in the 

sensitive endpoint of ALC was observed in the subgroup exposed to a median benzene level of 

7.6 ppm. Thus, this value is used as the LOAEL in Section 5.1.4. to derive a chronic inhalation 

RfC for benzene and in Section 5.2.4. to derive a chronic oral RfD. 

The primary quantitative analysis for developing the RfC and the RfD, however, uses 

BMD modeling approaches to model the Rothman et al. (1996a) ALC data and derive a BMD at 

the low end of the range of observation (Section 5.1.2). The BMDL is then used as the point of 

departure for the application of UFs in the calculation of the RfC or RfD (Sections 5.1.3 and 

5.2.3, respectively). Clearly, this is a preferable methodology because more of the exposureresponse data can be included in the analysis, and one is not restricted to selecting one of the 

exposure group medians as a LOAEL. The LOAEL-based RfC and RfD estimates are provided 

for comparison with the BMD-based estimates. RfC and RfD estimates are also derived from 

applicable experimental animal studies for comparison with the human-based estimates (Sections 

5.1.5 and 5.2.5, respectively). 

42
4.1.2.2. Neurotoxicity 

Humans have displayed symptoms of neurotoxicity following acute inhalation of 

relatively high concentrations of benzene (Snyder, 1987). Benzene produces generalized 

symptoms such as dizziness, headache, and vertigo at levels of 250–3000 ppm (799–9584 

mg/m3) (Brief et al., 1980), leading to drowsiness, tremor, delirium, and loss of consciousness at 

700–3000 ppm (2236–9584 mg/m3

) (ATSDR, 1997). Death may result from exposure to 20,000 

ppm (63,894 mg/m3

) benzene for 5–10 minutes (Sandmeyer, 1981). Neurological signs and 

symptoms are similar for both lethal and nonlethal exposures to benzene and for exposure to 

multiple solvents including benzene. These neurological symptoms are reversible upon removal 

of the subject from exposure (Kraut et al., 1988). 

Chronic exposure to benzene and toluene was studied in 121 workers exposed to benzene 

for 2–9 years (Kahn and Muzyka, 1973). The air concentration of benzene between 1962 and 

1965 was 6–15.6 ppm (20–50 mg/m3

), whereas toluene vapors did not exceed the 5 mg/m3 level. 

Subsequently, the air levels of both benzene and toluene did not exceed the 5 mg/m3 level. 

Seventy-four of the examined workers complained of frequent headaches (usually at the end of 

the workday), became tired easily, had difficulty sleeping, and complained of memory loss. The 

limitations of this study are that workers were exposed to both benzene and toluene, and the dose 

and duration of exposure were unknown. 

Few studies have examined the neurological effects of chronic exposure to benzene. 

Herregods et al. (1984) diagnosed transverse myelitis in a young man exposed daily to benzene 

while working in the warehouse of a wholesale supplier of drugs and chemicals. The duration of 

exposure was not given by the authors. The diagnosis of benzene poisoning was based on 

differential diagnosis, ruling out other possible causes of poisoning. The clinical diagnosis was 

based on the patient’s occupation, high urinary phenol levels (28 mg/L in contrast to the normal 

level, < 10 mg/L), and the coincidental decrease in urinary phenol and improvement of clinical 

symptoms over about 6 months after cessation of exposure. The diagnosis of transverse myelitis 

is consistent with an acute transection of the spinal cord affecting both the gray and white matter. 

Baslo and Aksoy (1982) conducted neurological examinations on eight patients who had 

a history of chronic exposure to benzene and who were diagnosed with aplastic anemia (cases 

1–6) or preleukemia (cases 7 and 8). The seven males and one female ranged in age from 19 to 

51 years (average age, 35.3 years). Cases 1, 2, 4, 5, 7, and 8 worked in the shoemaking industry, 

case 3 was a leather worker, and case 6 was a whistle maker who dipped plastic material into an 

open vessel of solution known to contain 88.42% benzene and 9.25% toluene. At times, the 

concentrations of benzene in the working environments reached 210 ppm (671 mg/m3

). 

Exposure duration for the six patients with aplastic anemia ranged from 6 months to 8 years 

(mean, 6 years), and the period between the cessation of exposure and the neurological 

examinations ranged from 1 to 96 months. The two patients who had preleukemia (cases 7 and 

8) were exposed for approximately 15 and 10 years, respectively, and were not exposed during 

the 8 and 6 months preceding neurological examination. Neurological, electromyographical 

(EMG), and motor conduction velocity examinations were performed on all patients (with the 

exception of case 8, who did not have the EMG test). At least three different muscles were 

tested in the EMG and motor conduction velocity examinations. Sensory conduction velocities 

were measured in the upper and lower extremities of cases 1, 5, 6, 7, and 8. 

43
Four cases (1, 2, 3, and 8) had other neurological symptoms that included decreased 

sensory vibration, atrophy of the leg, and exaggerated deep tendon reflexes; the other four (cases 

4, 5, 6, and 7) had normal neurological test results. EMG tests revealed neurogenic involvement 

in cases 1 and 2 but were normal in cases 3–7. Five of the cases (1, 2, 3, 7, and 8) had abnormal 

motor conduction velocity tests, involving at least one of the nerves tested. These effects were 

characterized by decreased motor conduction velocity and lengthening of latency in distal 

nerves. Conduction velocity was decreased in the sensory nerves of cases 1, 6, and 7. In case 6, 

the amplitudes of nerve action potentials were low. These effects reflected benzene-induced 

effects on the axons of the peripheral nerves. 

The investigators concluded that the neurological abnormalities in the four pancytopenic 

individuals resulted from a direct effect of benzene or toluene (case 6) on peripheral nerves 

and/or the spinal cord. They also concluded that the effects were related to the period of 

nonexposure. For example, cases 3 and 4 had moderate to severe hematologic findings and 

aplastic anemia, but no significant neurological findings. For these two cases, long periods of 

nonexposure (53 and 96 months, respectively) preceded neurological testing. 

The study is based on a small number of patients, lacks exposure and control data, and 

does not rule out the possibility that the workers were exposed to chemicals other than benzene 

and toluene; therefore, a reliable LOAEL or NOAEL could not be determined. The clinical 

impression indicated a predominant involvement of the white matter in the peripheral nervous 

system. 

The reports summarized in this section have obvious deficiencies, such as small numbers 

of patients under study and a lack of information regarding the intensity or duration of benzene 

exposure. Thus, a reliable quantitative neurotoxic risk assessment is impossible at present. 

Nevertheless, the occurrence of symptoms following exposure and the amelioration of effects 

upon removal from exposure leave little doubt that benzene affects the central as well as the 

peripheral nervous system. 

4.1.2.3. Reproductive Toxicity 

The available studies on the reproductive toxicity of benzene in humans are summarized 

in Table 4. 

Vara and Kinnunen (1946) examined 12 female workers who had gynecological 

disorders attributed to benzene exposure: three were 25–28 years old, seven were 37–40 years 

old, and two were 43–44 years old, and their exposure to benzene ranged from 1 to 10 years. 

The investigators identified the third group as “nearing menopause” but found no evidence of 

menopause; however, the symptoms of toxicity, consisting of hypermenorrhea and 

hypomenorrhea, ovarian hypoplasia, sterility, degeneration of the ovary, and/or dysfunction of 

the ovary were more severe in the two subjects nearing menopause. Peripheral blood counts 

revealed distinct leukopenia in four subjects and decreased neutrophils and platelets in most 

44
 Table 4. Reproductive toxicity of inhaled benzene in humans 

Population No. Exposure Effects Reference 

Female workers with 30 Overt symptoms suggested exposure to 12 subjects had menstrual disorders (e.g., hypermenorrhea) Vara and Kinnunen,

symptoms of benzene levels greater than 3.2 mg/m3 and/or ovarian hypoplasia; women nearing menopause most 1946

poisoning severely affected; 4 had leukopenia, and the majority had 

Female factory workers with benzene-induced ovarian hypofunction 

ND ND 

Female gluing operators 360 Exposure to gasoline (a major source of 

benzene) and chlorinated hydrocarbons via skin and inhalation; benzene levels, < 5 mg/m3; 40% of chlorinated 

hydrocarbon measurements exceeded permissible limits by 1.2- to 2.4-fold; controls had no chemical exposure 

Female solvent workers 174 (1) < 131 mg/m3 benzene 

(2) 131 mg/m3 benzene 

(3) control 

Female workers in shoemaking industry 

223 Subjects exposed to benzene and toluene 

(compared with 327 controls) 

reduced neutrophil and platelet counts 

Workers exhibiting ovarian hypofunction had decreased 

ascorbic acid concentrations in the blood, as compared with clinically healthy female workers (p<0.001) 

Functional disturbances of the menstrual cycle 

(1) Hypermenorrhea in 4/40 (10%) (p<0.05) (2) Hypermenorrhea in 4/47 (8.5%) (p<0.05) (3) Hypermenorrhea in 1/87 (1.2%) Increased menstrual disorders (exposed, 48.9%; control, 16.2%); spontaneous abortion (exposed, 5.7%; control, 2.4%); gestosis (exposed 22.6%; control, 10.5%); all differences were statistically significant 

Pushkina et al., 1968 

Mukhametova and 

Vozovaya, 1972 

Yin et al., 1987a Huang., 1991 

ND = no data. 

45
subjects. The investigators tentatively attributed the sparseness in menstruation to the ovarian 

hypoplasia rather than to the benzene. 

Pushkina et al. (1968) evaluated a group of workers who had ovarian hypofunction 

related to exposure to benzene. Compared with clinically healthy female workers in the same 

factory, the levels of ascorbic acid (changes in ascorbic acid metabolism were used as an index 

of toxicity) were reduced in the blood of the workers who had ovarian dysfunction (subjects, 

0.36 ± 0.02 mg%; controls, 0.49 ± 0.03 mg%; p<0.001). The details of this study were sparse, 

but the results were supported by a study in rats that showed a similar effect. The justification for 

the changes in the ascorbic acid as an index of toxicity was not well desribed. 

A Russian study evaluated 360 women exposed to gasoline (a source of exposure to 

benzene) and chlorinated hydrocarbons via inhalation and skin contact (Mukhametova and 

Vozovaya, 1972). Benzene levels were < 1.6 ppm (5 mg/m3

); chlorinated hydrocarbon levels 

were 1.2–2.4 times higher than the permissible limits. Compared with workers with no exposure 

to benzene, the subjects had increased incidences of menstrual disturbances. As exposure 

duration increased, so did the number of premature interruptions of pregnancy, the percentage of 

cases where theplacental membrane rupture during parturition was impeded, and the number of 

cases of intrauterine asphyxia of the fetus. This study, although demonstrating exposure-related 

reproductive effects and fetal toxicity, fails to distinguish between the effects of benzene and 

chlorinated hydrocarbons. 

Yin et al. (1987a) conducted a comprehensive study of the toxicity of benzene in which 

300 workers exposed to benzene, toluene, or a mixture of the two were examined for subjective 

symptoms and hematologic and biochemical effects. For 174 women, the mean TWA exposure 

to benzene was 59 ppm (188 mg/m3

) (exposure duration, 65 months). The women were divided 

into low-exposure (1–40 ppm [3.2–128 mg/m3

]) and high-exposure (41–210 ppm [131–671 

mg/m3

]) groups and were examined separately for subjective and pancytopenia-related subjective 

symptoms. The benzene-exposed subjects reported an increase in the incidence of 

hypermenorrhea, but the effect was not concentration-related (prevalence was 10% for lower 

exposures and 8.5% for higher exposures; p<0.05 for both doses). Although the only 

hematologic abnormality noted was a mild but statistically significant decrease in lymphocyte 

count (p<0.05), this hypermenorrhea was tentatively considered to be related to pancytopenia. 

These findings are strengthened by the large number of subjects, the measured exposure and 

statistical evaluation, and similar observations in other studies; however, the significance of 

subjective symptoms is debatable. 

A Chinese report (Huang, 1991) described reproductive dysfunction in female workers 

exposed to benzene and toluene in the leather shoemaking industry. The exposed group (223 

women) exhibited increased incidence in the rate of menstrual disorders as compared with the 

327 controls. The incidence rates of “mense-blood anomaly” and dysmenorrhea had a tendency 

to increase with duration of employment. The incidences of spontaneous abortion and gestosis 

(toxemia) were also increased. All increases were statistically significant. The investigators 

concluded that both benzene and toluene had a deleterious effect on the reproductive function of 

female workers. 

46
The data presented above provide limited evidence regarding reproductive effects of 

benzene exposure in humans and involve concomitant exposure to other chemicals. These 

studies did not provide good exposure monitoring data or quantitative dose-response 

information. 

4.1.2.4. Developmental Toxicity 

The following discussion reviews the developmental toxicity of benzene in humans, 

focusing on more recent reports. Available studies on the developmental toxicity of benzene in 

humans are summarized in Table 5. There is no convincing evidence that benzene produces 

malformations in humans or test animals; however, a few studies suggest that the chemical 

induces adverse reproductive effects in humans and causes retarded fetal growth in test animals, 

manifested mainly as decreased fetal weight and delayed ossification in the presence or absence 

of maternal toxicity (Chatburn et al., 1981; ATSDR, 1997). 

Benzene crosses the human placenta and is present in cord blood in quantities equal to or 

greater than those in the maternal blood (Dowty et al., 1976). Summaries of developmental 

toxicity of benzene in humans have been described in individual case reports (Forni et al., 1971; 

Holmberg, 1979; Bordarier et al., 1991) and occupational studies (Vara and Kinnunen, 1946; 

Pushkina et al., 1968; Mukhametova and Vozovaya, 1972; Funes-Cravioto et al., 1977; Axelsson 

et al., 1984; Yin et al., 1987a; Savitz et al., 1989; Huang, 1991). Exposures were either to 

benzene alone or, as is characteristic of occupational exposure, to multiple chemicals including 

benzene. With only two exceptions (Bordarier et al., 1991; Mukhametova and Vozovaya, 1972), 

all adult exposures in the studies summarized in these sections were via inhalation. 

Only one study evaluated fetal effects following inhalation of benzene alone. A worker 

who was exposed to benzene during her entire pregnancy suffered from severe pancytopenia and 

exhibited increased chromosomal aberrations, but there were no effects on the fetus (Forni et al., 

1971). 

Axelsson et al. (1984) evaluated by questionnaire the outcome of pregnancy among 

personnel employed in laboratory work at the University of Gothenburg between 1968 and 1979. 

Of 745 women who responded to the questionnaire, 556 had been pregnant (a total of 1160 

pregnancies). The pregnant women were divided into two groups: those with exposure to 

organic solvents during laboratory work and those without exposure. Responders to the 

questionnaire reported exposure to at least 14 solvents; 41 workers remembered using benzene 

during the first trimester of pregnancy and 5 used phenol. A slightly increased but not 

significant difference in the miscarriage rate was found (relative risk = 1.31, 95% CI = 

0.89–1.91). There were no differences in perinatal death rates or prevalence of malformations 

between infants 

whose mothers were exposed to solvents and those who were not exposed. The investigators 

suggested that shift work, number of pregnancies, and age may have contributed to an increase in 

miscarriage rate in the group of pregnant women not exposed to solvents, resulting in an 

underestimation of the miscarriage rate of the exposed women. An additional concern was that 

in spite of the fact that confounders were reduced by using as controls laboratory workers not 

exposed to solvents, a mutagenic effect occurring during laboratory work (laboratory work the 

woman had stopped working at the laboratory, thereby altering the miscarriage rate in the 

47 
Table 5. Developmental toxicity of benzene—humans 

Population No. Exposure Effects Reference 

Pregnant worker 1 Exposure to benzene during entire Maternal effects included severe pancytopenia and increased Forni et al., 1971 

chromosomal aberrations; no fetal effects pregnancy 

Employees in university laboratory (~745 subjects, 1160 pregnancies), divided into those with and without exposure to organic solvents 

745 Responders to questionnaire reported 

exposure to at least 14 solvents; 41 workers remembered using benzene 

during the first trimester of pregnancy and 5 used phenol 

All women exposed to solvents had slight but not statistically significant difference in miscarriage rate over those not exposed (RR = 1.31, 95% CI = 0.89–1.91); 35 of 41 workers exposed to benzene delivered, 1 had induced abortion, and 5 miscarried (miscarriage rate for benzene-exposed subjects, 12.2%; miscarriage 

rate for all responders to questionnaire, 11.1%; miscarriage rate for 

unexposed responders, 10.1%); all 5 workers exposed to phenol delivered; exposure to solvents did not affect perinatal death rates or the incidence of malformations 

Axelsson et al., 1984 

Female gluing operators 360 Exposure to gasoline (a major source of 

benzene) and chlorinated hydrocarbons via skin and inhalation; benzene levels, < 5 mg/m3; 40% of chlorinated 

hydrocarbon measurements exceeded permissible limits by 1.2- to 2.4-fold; controls had no chemical exposure 

Spontaneous abortions and premature births (17.2% vs. 4.9% in controls), incidence of late membrane rupture, and intrauterine asphyxia of the fetus increased with exposure duration 

Mukhametova and Vozovaya, 1972 

Adult female workers 

and 14 of their children 

29 Adults exposed to benzene and other 

organic solvents during pregnancy (compared with 42 control adults and 7 of their children) 

Lymphocytes from adults exhibited approximately twofold increase over controls in incidence of chromosomal aberrations and breaks; their children exhibited increased frequency of chromatid breaks, isochromatid breaks (p<0.01, 14 children), and sister chromatid exchanges (p<0.001, 4 children) in lymphocytes 

Funes-Cravioto et al., 1977 

32-year-old pregnant worker 

1 Personal interview revealed exposure to 

benzene and other solvents (dichloromethane, methanol, and ether) in laboratory during first trimester of pregnancy; compared with matched control 

Stillborn anencephalic fetus Holmberg, 1979 

48
Table 5. Developmental toxicity of benzene—humans (continued) 

Reference 

Bordarier et al., 1991 

Effects 

Following normal delivery, infant exhibited slight dysmorphy 

(hypotelorism and deep nasal bridge), moderate axial hypotonia and abnormal ocular movements; at 1.5 months of age, child was microcephalic, had severe axial hypotonia, severe peripheral 

hypertonia, and bilateral optic atrophy, and CT scanning revealed 

bilateral porencephalic cavities that created communication between lateral ventricles and subarachnoid space; interventricular septum lacking; child died from aspiration pneumonia at 2 months of age 

Exposure 

21 intramuscular injections of benzene in 

an unsuccessful attempt to induce abortion during first trimester of pregnancy 

No. 

1 

Population 

23-year-old female 

RR = relative risk. CT = computerized transverse tomography. 

49
control group. Another weakness in the study was that the subjects had worked in the laboratory 

during 1968–1979, and some had to remember several years back. 

Four other studies evaluated the developmental toxicity of multiple solvents. These 

effects consisted of spontaneous abortions, premature births, and effects on “condition of the 

fetus” (Mukhametova and Vozovaya, 1972), chromosomal abnormalities (Funes-Cravioto et al., 

1977), and stillbirth (Holmberg, 1979). 

Savitz et al. (1989) examined the effects of parental (paternal as well as maternal) 

occupational exposures on fetal development. The subjects, employed in various industries, 

were exposed to a number of chemicals. Summarized here are the data pertaining only to 

benzene exposures. The investigators accessed the National Natality and Fetal Mortality surveys 

to obtain data on probability samples of live births and fetal deaths that occurred in the United 

States in 1980 among married women. These data were merged with data provided by mothers 

from questionnaires and from information from medical care providers for the public-use data 

set. Omitted from the analysis were unmarried mothers, mothers who had not worked within 12 

months of delivery, plural births, and births of unknown plurality. 

The analysis, which was based on three case control studies of pregnancy outcome, 

included case groups of stillbirths, preterm deliveries (birth before 37 weeks of gestation), and 

small-for-gestational-age infants. The investigators used an exposure linkage system to 

designate exposures. The exposure to a specific agent was assigned on the basis of the 

occupation and industry of employment and on past studies and review of industrial processes. 

Linkages of none, low, medium, high, or unknown were assigned to each agent, representing 

probability (rather than intensity) of exposure. Unexposed working women served as controls. 

The odds ratios (ORs) relating exposure to effect were adjusted for selected potential 

confounders. Elevated ORs were found for maternal exposure to benzene and stillbirth (OR = 

1.3, 95% CI = 1.0–1.8). Increased risks across low-, medium-, and high-linkage levels (crude 

ORs of 0.9, 1.2, and 1.4, respectively) were observed. Exposure to benzene was also associated 

with risk elevation for the fathers (OR = 1.5, 95% CI = 1.1–2.3). The OR for low, medium, and 

high linkages were 1.2, 1.5, and 2.0, respectively, with reasonably good precision for the two 

highest levels (95% CI = 1.0–2.2 and 1.1–3.7, respectively). Benzene exposure linkage level 

was unrelated to risk of preterm delivery for both parents. The investigators recognized that the 

study had limitations. These included small population sizes, poor quality of exposure 

information, and absence of statistical testing and the high rate of nonresponse to questionnaires 

among women who were under 20 years old, 40 or more years old, and black and who had parity 

four or greater, little or no prenatal care, and low education. However, in spite of these 

limitations, the authors concluded that the results of the study encourage further evaluation of the 

developmental effects of paternal exposure to benzene. 

The data presented above provide inconclusive results regarding developmental toxicity 

of benzene in humans. Most studies consisted of small numbers of subjects, lacked important 

experimental details, and involved (in almost all cases) concomitant exposure to other chemicals, 

and they did not provide monitoring data or quantitative dose-response information. 

4.2. ACUTE AND CHRONIC STUDIES IN EXPERIMENTAL ANIMALS 

50 
4.2.1. Hematotoxicity 

4.2.1.1. Oral Exposure—Subchronic Studies 

Subchronic oral studies have been conducted in F344 rats and B6C3F1 mice of both 

sexes (NTP, 1986; Huff et al., 1989) and female Wistar rats (Wolf et al., 1956). Other 

subchronic oral studies have been conducted to examine immunotoxicity, and these are 

described in Section 4.2.4.1. 

Hsieh et al. (1988b) exposed male CD-1 mice to benzene in drinking water at 

concentration levels 0, 31, 166, or 790 mg/L (0, 8, 40, or 180 mg/kg/day) for 28 days. Doserelated hematologic effects (erythrocytopenia, leukocytopenia, lymphocytopenia, increased 

MCV) were observed at all exposure levels. It can be estimated that these dose levels in 

drinking water are roughly equivalent to inhaled doses of 6.4, 32, and 154 ppm. A major 

drawback of this study is that there were only five animals per group. Details of the study are 

discussed in Section 4.2.4.1. 

In the National Toxicology Program (NTP) study (NTP, 1986), F344 rats and B6C3F1 

mice (10/species/group/sex; 6–8 weeks of age) were treated with 0, 25, 50, 100, 200, 400, or 600 

mg/kg benzene by gavage in corn oil 5 days/week for 17 weeks. The adjusted doses were 0, 18, 

36, 71, 143, 286, or 429 mg/kg/day. An additional five animals/species/group/sex were tested at 

the 0, 200, or 600 mg/kg dose levels and killed at 60 days of treatment. Hematologic analyses 

were performed on all the animals killed at 60 days and on five animals/species/group/sex at the 

end of the study. In addition, necropsies were performed on all animals, and the spleens of all 

animals were examined histopathologically. 

No compound-related deaths were observed for rats. Final body weight depression of 

10% relative to controls was observed in male and female rats at dose levels of 200 mg/kg and 

greater. Significant (p<0.05) leukopenia and lymphocytopenia were observed in male and 

female rats after 60 days of treatment with 200 or 600 mg/kg (the only treatment groups tested 

on day 60). On day 120 of treatment, significant leukopenia and lymphocytopenia were 

observed in female rats at 25 mg/kg and higher, and significant lymphocytopenia was observed 

in male rats at 400 mg/kg. Lymphoid depletion of B-cells in the spleen was observed in 100% of 

male and female rats exposed to 600 mg/kg for 60 or 120 days. Increased extramedullary 

hematopoiesis in the spleen was observed in four of five male and three of five female rats 

treated with 600 mg/kg for 120 days. This study identified a LOAEL of 25 mg/kg (18 

mg/kg/day) in female rats and a LOAEL of 200 mg/kg (143 mg/kg/day) in male rats for 

hematologic effects following treatment by gavage for 17 weeks. The LOAEL for female rats 

was at the lowest dose tested; thus, no NOAEL was established. 

There were no compound-related deaths in the mice. A final body weight depression of 

7% was seen in the 100 mg/kg dose group. Tremors were observed intermittently in male and 

female mice treated with 400 or 600 mg/kg. No leukopenia or lymphocytopenia was observed in 

male or female mice after 60 days of treatment with 200 or 600 mg/kg. At 120 days, significant 

(p<0.05) leukopenia and lymphocytopenia were observed in male mice at dose levels of 50 

mg/kg and greater and in female mice at 400 (only lymphocytopenia) and 600 mg/kg. A 

51 
NOAEL of 25 mg/kg (18 mg/kg/day) and a LOAEL of 50 mg/kg (36 mg/kg/day) for 

hematologic effects were identified in male mice treated by gavage for 17 weeks. A NOAEL of 

200 mg/kg (143 mg/kg/day) and a LOAEL of 400 mg/kg (286 mg/kg/day) for hematologic 

effects were identified in female mice treated by gavage for 17 weeks. 

Female Wistar rats (10/group) were treated by gavage with benzene in olive oil 5 

days/week for 6 months (Wolf et al., 1956). The reported doses were 0, 1, 10, 50, or 100 mg/kg 

(0, 0.7, 7.1, 35.7, and 71.4 mg/kg/day). Parameters measured included mortality, clinical signs, 

body and organ weights, hematology, blood biochemistry, bone marrow counts, and gross and 

microscopic pathology of lungs, heart, liver, kidneys, spleen, testes, adrenals, and pancreas. 

Leukopenia was reported for 10 mg/kg; at higher dose levels erythrocytopenia and leukopenia 

were also observed. No quantitative data or statistical analysis were reported. The authors 

reported that rats fed 1 mg/kg had “no evidence of ill effects.” This study identified a NOAEL 

of 0.7 mg/kg/day and a LOAEL of 7.1 mg/kg/day for hematologic effects in female rats treated 

by gavage for 6 months. 

Aoyama (1986) showed that a 14-day exposure of mice to 50 ppm (162 mg/m3

) benzene 

resulted in a significantly reduced blood leukocyte count. 

4.2.1.2. Oral Exposure—Chronic Studies 

In the NTP (1986) study, F344 rats and B6C3F1 mice of both sexes were treated by 

gavage with benzene 5 days/week for 103 weeks. Results of this study have also been reported 

by Huff et al. (1989). For rats, males (60/group) were administered doses of 0, 50, 100, or 200 

mg/kg (0, 36, 71, or 143 mg/kg/day), and females (60/group) were administered doses of 0, 25, 

50, or 100 mg/kg (0, 18, 36, or 71 mg/kg/day). Survival decreased with increasing dose in rats 

of both sexes and was significantly decreased (p<0.05) at 200 mg/kg in males and at 50 and 100 

mg/kg in females. Body weight depression of 10% relative to controls was observed in male 

rats treated with 100 mg/kg. Dose-related leukopenia was significant in female rats treated with 

25 mg/kg for 3, 6, 9, and 12 months; leukocyte levels were comparable with controls after 15, 

18, 21, and 24 months of treatment. In male rats, dose-related leukopenia was significant at 50 

mg/kg. Lymphoid depletion was observed in the thymus of 0/44, 4/42, 8/41, and 10/34 male rats 

treated with 0, 50, 100, and 200 mg/kg benzene, respectively. In the spleen, lymphoid depletion 

was observed in 0/49, 19/58, 8/47, and 23/47 male rats treated with 0, 50, 100, and 200 mg/kg, 

respectively. This study identified a LOAEL of 25 mg/kg (18 mg/kg/day) for leukopenia and 

lymphocytopenia in female F344 rats and 50 mg/kg (36 mg/kg/day) in male F344 rats. These 

were the lowest doses tested, and thus no NOAEL was identified. 

B6C3F1 mice (60/sex/group) were administered doses of 0, 25, 50, or 100 mg/kg (0, 18, 

36, or 71 mg/kg/day). Survival decreased with increasing dose in mice of both sexes and was 

significantly decreased (p<0.05) at 100 mg/kg. Body weight depression of 10% relative to 

controls was observed in mice of both sexes treated with 100 mg/kg. Significantly decreased 

leukocyte counts were observed in males after 3, 6, 9, 12, 15, 18, and 21 months of treatment 

with 50 or 100 mg/kg, but males treated with 25 mg/kg had significantly decreased leukocyte 

counts only after 6 and 21 months. In female mice, leukopenia was observed only at 12 and 18 

months, in both cases significant at all treatment levels. Significantly decreased lymphocyte 

counts were observed in males after 3, 6, 9, 12, 15, 18, and 21 months with 50 or 100 mg/kg, but 

52 
males treated with 25 mg/kg had significantly decreased lymphocyte counts only after 12 

months. In female mice, significant lymphocytopenia was observed at 25 mg/kg at 12 and 18 

months and at 100 mg/kg at 3 months. Hematopoietic hyperplasia of the bone marrow was 

observed in 0/49, 11/48, 10/50, and 25/49 male mice and in 3/49, 14/45, 8/50, and 13/49 female 

mice treated with 0, 25, 50, or 100 mg/kg, respectively. Increased splenic hematopoiesis was 

observed in 5/49, 9/48, 19/49, and 24/47 male mice and in 9/49, 10/45, 6/50, and 14/49 female 

mice treated with 0, 25, 50, or 100 mg/kg, respectively. In the female mice, increased incidences 

of epithelial hyperplasia of the ovary occurred at all three doses, and increased incidence of 

senile atrophy of the ovary occurred at the lower two doses compared with controls. This study 

identified a LOAEL of 25 mg/kg (18 mg/kg/day) for leukopenia and lymphocytopenia in male 

and female B6C3F1 mice. The observed LOAEL was at the lowest dose tested; thus, a NOAEL 

was not identified. 

Beginning in 1976, a series of carcinogenicity studies on oral treatment of rodents with 

benzene were performed at the Bologna Institute of Oncology, including 52- and 104-week 

studies on Sprague-Dawley and Wistar rats and Swiss and Rf/J mice (Maltoni et al., 1983, 1985). 

Limited information regarding noncarcinogenic effects were reported for Sprague-Dawley rats, 

but only carcinogenicity data were published for Wistar rats and mice. No statistical information 

was included, making interpretation of the data difficult. 

Maltoni et al. (1985) treated Sprague-Dawley rats (13 weeks of age, 30–35/sex/group) by 

gavage with 0, 50, or 250 mg/kg benzene in oil 4–5 days/week for 52 weeks and then observed 

them until death. Expanded doses were 0, 32, and 161 mg/kg/day. In addition, Sprague-Dawley 

rats (7 weeks of age, 40–50/sex/group) were treated by gavage with 0 or 500 mg/kg benzene in 

oil 4–5 days/week for 104 weeks and then observed until death. The expanded doses were 0 and 

321 mg/kg/day. Mortality was higher in benzene-treated groups and appeared to be dose related; 

body weights were not affected. Maltoni et al. (1983) stated that mortality in the first portion of 

the study was due to direct toxic effects of treatment, and in the later portion it was partially due 

to tumors. Mortality was similar to that of controls during treatments with 500 mg/kg for 92 

weeks; body weight appeared to be somewhat depressed relative to controls. In Sprague-Dawley 

rats exposed to 500 mg/kg for 84 or 92 weeks, decreased total RBCs (only at 92 weeks), WBCs, 

and lymphocytes were observed. Insufficient information was provided to establish LOAEL or 

NOAEL levels from these studies. 

4.2.1.3. Inhalation Exposure—Subchronic Studies 

The following discussion summarizes more recent studies of the effects of inhaled 

benzene on the peripheral blood of experimental animals. The available studies on the 

hematotoxicity of inhaled benzene to test animals are summarized in Table 6. These studies 

form a database that supports the understanding that exposure to benzene at high concentrations 

causes hematotoxic effects, including bone marrow suppression and the greater sensitivity of 

mice over that of rats. Benzene-induced hematotoxicity has been observed from short-term as 

well as long-term exposure to the chemical. 

53
Table 6. Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposure 

Strain/Species Endpoints Exposure Effects 

LOAEL/NOAEL 

(mg/m3) Reference 

CD-1 mice (11–12 males/ 

group) 

Body weight, cells in peripheral blood and bone marrow, spleen weight, total nucleated cells per spleen, lymphocytes 

0, 32 mg/m3, 6 

hr/day, 5 days/wk for 10 weeks 

Increased spleen weight, total nucleated cells per spleen, and nucleated cells per spleen 

LOAEL: 32 Green et al., 

1981a, b 

CD-1 mice 

(30 males and 30 females/group) 

HCT; Hgb; WBC, RBC, platelet, reticulocyte, and differential counts; MCV; mean corpuscular Hgb; myeloid/erythroid ratio from bone marrow; leukocyte alkaline phosphatase; RBC glycerol lysis time; and other more general parameters of subchronic toxicity 

0, 3.2, 32, 96, or 958 mg/m3 

6 hr/day, 5 days/wk for up to 91 days; serial sacrifices after 7, 14, 28, 56, and 91 days of exposure 

At 958 mg/m3, decreased HCT, total 

Hgb, RBC, leukocyte, and platelet counts; decreased myeloid/erythroid ratios and percentage of lymphocytes; increased MCV, mean cell Hgb, glycerol lysis time, and incidence and severity of RBC morphological changes (statistically significant [p<0.05] for males and/or females); many effects occurred and were statistically significant by 14 days, persisting throughout exposure; microscopic changes in the thymus, bone marrow, lymph nodes, spleen, ovaries, and testes; changes in males occurred more often and with greater severity than those in the females 

LOAEL: 958 NOAEL: 96 

Ward et al., 1985 

AKR/J mice 

(60 males, test; 60 males, control) 

WBC and RBC counts (Coulter), reticulocyte counts, differential counts, Hgb, HCT, lactic acid dehydrogenase, RBC acetylcholinesterase, reduced GSH, cytogenetic analyses of bone marrow, and 

microscopic examination of lung, liver, spleen, and kidney 

958 mg/m3 

benzene, 6 hr/day, 5 days/wk for life 

Early mortality (mean survival time, 11 weeks for test animals, 39 weeks for controls), severe lymphocytopenia, bone marrow hypoplasia, granulocytosis, and reticulocytosis 

LOAEL: ND NOAEL: ND 

Snyder et al., 1978 

54
Table 6. Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposure (continued) 

Reference 

Snyder et al., 1980 

Snyder et al., 1980 

LOAEL/NOAEL 

(mg/m3) 

LOAEL: ND NOAEL: ND 

LOAEL: ND NOAEL: ND 

Effects 

No statistically significant differences between test and control mice in median survival time, rate of weight gain or lymphoma type or occurrence; 

statistically significant lymphocytopenia (cell counts were depressed to ~65% of normal throughout exposure); depressed RBC counts (± 2 standard error of mean at 9/19 monitoring periods); increased neutrophils (± 2 standard error of mean at 3/19 monitoring periods); bone marrow hypoplasia in 20% of the treated mice (p<0.05) 

Decreased survival (median time 41 weeks, test animals; 75 weeks, controls); lymphocytopenia (WBC 

counts, ~3000/mm3 or 15% of normal); 

anemia (RBC counts, 8 × 106/mm3 or 

~80% of normal); neutrophilia appearing at 17 weeks; abnormal blood cell morphology; neutrophilic shift to left; bone marrow hyperplasia in 33% of test animals, none in controls; hematopoietic neoplasms 

Exposure 

319 mg/m3 

benzene, 6 hr/day, 5 days/wk for life 

958 mg/m3 

benzene, 6 hr/day, 5 days/wk for life 

Endpoints 

RBC and WBC counts (Coulter) and differential count; microscopic examination of lung, liver, spleen, kidney, and bone marrow 

RBC and WBC counts (Coulter) and differential count; microscopic examination of lung, liver, spleen, kidney, and bone marrow 

Strain/Species 

AKR/J mice (50 males, test; 50 males, control) 

C57BL/6 mice (40 males, test; 40 males, control) 

55
Table 6. Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposure (continued) 

Reference 

Snyder et al., 1982 

Ward et al., 1985 

LOAEL/NOAEL 

(mg/m3) 

LOAEL: ND NOAEL: ND 

LOAEL: 958 NOAEL: 96 

Effects 

Decreased median survival time (25.5 weeks for benzene-exposed animals, 52.7 weeks for controls); depressed peripheral RBC and lymphocyte counts (± 2 standard error of mean); neutrophilia; abnormal blood cell morphology; and a shift to immature myeloid cells (at 217 days); tumor incidence (5/40) not significant; acute myeloblastic leukemia in one exposed animal, chronic myelogenous leukemia in one; bone marrow hyperplasia in 9/35 animals without neoplasia, bone marrow hypoplasia in 11, splenic hemosiderin pigments in 6, and splenic hyperplasia in 19 Statistically significant (p<0.05) decrease in WBC counts (males on day 14 and females on days 14–91) and slightly decreased femoral marrow cellularity at 958 mg/m3 (300 ppm) 

Exposure 

958 mg/m3 

benzene, 6 hr/day, 5 days/wk for life 

0, 3.2, 32, 96, or 958 mg/m3 

6 hr/day, 5 days/wk for up to 91 days; serial sacrifices after 7, 14, 28, 56, and 91 days of exposure 

Endpoints 

RBC and WBC counts (Coulter), and differential count; microscopic examination of lung, liver, spleen, kidney, and bone marrow 

HCT; Hgb; WBC, RBC, platelet, reticulocyte, and differential counts; MCV; mean corpuscular hemoglobin; 

myeloid/erythroid ratio from bone marrow; leukocyte alkaline phosphatase; erythrocyte glycerol lysis time; and other more general parameters of subchronic toxicity 

Strain/Species 

CD-1 mice (40 males, test; 40 males, control) 

SpragueDawley rats (10 males and 10 females/group) 

56
Table 6. Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposure (continued) 

Strain/Species Endpoints Exposure Effects 

LOAEL/NOAEL 

(mg/m3) Reference 

SpragueDawley rats (45 males, test; 

27 males, control) 

WBC and RBC counts (Coulter), reticulocyte counts, differential counts, Hgb, HCT, lactic acid dehydrogenase, RBC acetylcholinesterase, and reduced GSH, cytogenetic analyses of bone marrow, and 

microscopic examination of lung, liver, spleen, and kidney 

958 mg/m3 

benzene, 6 hr/day, 5 days/wk for life 

Mild weight depression, lymphocytopenia, trend to anemia, fatty changes in 77% of bone marrow samples 

LOAEL: ND NOAEL: ND 

Snyder et al., 1978 

Female BDF1 

mice (4/group) 

WBCs, reticuloctyes, differential counts, and assays for CFU-S, BFU-E, CFU-E, and GM-CFU-C 

0, 319, 958, or 2875 mg/m3, 6 

hr/day, 5 days/wk for 16 weeks 

Reduced lymphocyte count at 319 

mg/m3, decreased CFU-E, BFU-E, 

CFU-S, and GM-CFU-C 

LOAEL: 319 NOAEL: NA 

Seidel et al., 1989 

Male C57BL/6J mice (5/group) 

Peripheral blood counts, spleen and bone marrow BFU-E and 

CFU-E 

32 mg/m3, 6 

hr/day, 5 days/wk for up to 178 days 

Reduced circulating RBCs, lymphocyte 

counts, bone marrow CFU-E and BFUE, and splenic CFU-E 

LOAEL: 32 Baarson et al., 

1984 

Male and female C57/6 

BNL mice (5–10/group) 

WBC, RBC, differential counts, HCT, bone marrow 

cellularity, and CFU-S in bone marrow and spleen 

0, 32, 80, or 1278 mg/m3, 6 

hr/day, 5 days/wk for 2 weeks 

Reduced lymphocyte counts after 2 weeks of exposure at 80 mg/m3 but no 

effect at 32 mg/m3 

LOAEL: 80 NOAEL: 32 

Cronkite et al., 1985 

Male SpragueDawley rats Erythrocyte and lymphocyte counts 0 or 319 mg/m3, 6 

hr/day, 5 days/wk for life 

No significant reductions in erythrocyte or lymphocyte counts 

NOAEL: 319 American 

Petroleum Institute, 1983 

57
Table 6. Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposure (continued) 

Strain/Species Endpoints Exposure Effects 

LOAEL/NOAEL 

(mg/m3) Reference 

Male B6C3F1 mice (24/group exposed, bone marrow tests on 3–16/group) 

Bone marrow; cell counts, CFU-E, GM-CFU-C, and Band T-lymphocytes 

0, 3.2, 32, 319, or 639 mg/m3, 6 

hr/day, 5 days/wk for 1, 2, 4, or 8 weeks 

Reduced bone marrow cellularity, progenitor cells, and differentiating hematopoietic cells, at 319 mg/m3, but 

no effects observed at 32 mg/m3 

LOAEL: 319 NOAEL: 32 

Farris et al., 1996, 1997a 

Male B6C3F1 mice (3 or 9/group) 

Splenic, thymic, and femoral 

lymphocytes and labeling index of femoral lymphocytes 

0, 3.2, 16, 32, 319, or 639 mg/m3, 6 

hr/day, 5 days/wk for 1, 2, 4, or 8 weeks 

Persistent and rapid reductions in splenic and femoral lymphocytes at 

319 mg/m3, transient reduction of 

thymic cell count and splenic Blymphocytes at 32 mg/m3 after 2 weeks, 

but comparable with controls by 4 weeks 

LOAEL: 319 NOAEL: 32 

Farris et al., 1997b 

Male SpragueDawley rats (16/group) 

Thymus and spleen weight, spleen and bone marrow 

cellularity, spleen CD4+/CD5+ , 

CD8-/CD5+ and Kappa+ 

lymphocytes 

0, 96, 639, or 1278 mg/m3, 6 

hr/day, 5 days/wk for 2 or 4 weeks 

Reduced spleen B-CD4+/CD5+ and CD5+ T-lymphocytes at 1278 mg/m3 , 

no effects observed at 639 mg/m3 

LOAEL: 1278 NOAEL: 639 

Robinson et al., 1997 

ND = not determined because only one concentration was used.NA = not applicable, adverse effects were observed at the lowest dose tested.CFU-S = colony-forming unit–spleen.BFU-E = burst-forming unit–erythroid.CFU-E = colony-forming unit–erythroid.GM-CFU-C = granulocyte/macrophage colony-forming unit–culture

58
Early general toxicity studies reported leukopenia in dogs and fatal anemia in mice 

exposed to 600 ppm (1917 mg/m3

) benzene for 12–15 days (Hough and Freeman, 1944); changes 

in bone marrow histopathology or leukopenia in rats, guinea pigs, and rabbits exposed to 80–85 

ppm (256–272 mg/m3

) benzene for 23–187 exposures (Wolf et al., 1956); and leukopenia in rats 

exposed to 61 ppm (195 mg/m3

) benzene for 2–4 weeks or to 44 ppm (141 mg/m3) for 5–8 weeks 

(Deichmann et al., 1963). 

Male CD-1 mice (11–12/group) were exposed for 6 hours/day, 5 days/week to 

concentrations of 0 or 10 ppm (0 or 32 mg/m3

) benzene for 10 weeks or to 0 or 300 ppm (0 or 

958 mg/m3

) for 26 weeks (Green et al., 1981a,b). On the day of the last exposure, samples 

(pooled from groups of 3–4 mice) were obtained from the peripheral blood, bone marrow, and 

spleen to evaluate hematologic and hematopoietic cells. In mice exposed to 10 ppm (32 mg/m3), 

no adverse effects were observed with respect to mortality, body weight, or cells in the 

peripheral blood or bone marrow. Spleen weight, total nucleated cells per spleen, and nucleated 

RBCs per spleen were significantly increased (p<0.05) in mice exposed to 10 ppm (32 mg/m3

). 

Mice exposed to 300 ppm (958 mg/m3

) had the following significant (p<0.05) changes: 

increased mortality rate; decreased numbers of lymphocytes and RBCs in peripheral blood; 

decreased granulocyte/macrophage progenitor cells in bone marrow; decreased spleen weight 

and numbers of lymphocytes; multipotential hematopoietic stem cells and committed 

granulocyte/macrophage progenitor cells in the spleen; and increased incidence of atypical cell 

morphology in the peripheral blood, bone marrow, and spleen. These studies identify a LOAEL 

of 10 ppm (32 mg/m3

) for slight hematopoietic effects in mice exposed to benzene for 10 weeks. 

In a subchronic inhalation toxicity study, Ward et al. (1985) evaluated the peripheral 

blood and bone marrow of benzene-exposed CD-1 mice and Sprague-Dawley rats. Groups of 50 

male and 50 female rats and 150 male and 150 female mice inhaled benzene concentrations of 0, 

1, 10, 30, or 300 ppm (0, 3.2, 32, 96, or 958 mg/m3) 6 hours/day, 5 days/week for up to 13 

weeks. Serial sacrifice of 30 mice and 10 rats/group took place after 7, 14, 28, 56, and 91 days 

of exposure. The parameters of toxicity evaluated included behavior, body weight, organ 

weights, clinical pathology, gross pathology, and histopathology. Chamber analyses of benzene 

concentrations calibrated daily by gas chromatography and monitored continuously by infrared 

analyzer were within 10% of the target concentrations. 

Clinical observations and body weight data revealed no signs of exposure-related 

toxicity. Hematologic effects were not observed in either species at 1, 10, or 30 ppm (3.2, 32, or 

96 mg/m3

). The mice exposed to 300 ppm (958 mg/m3) for 91 days, however, exhibited 

decreased HCT and reductions in total Hgb concentration, RBC count, WBC count, platelet 

count, myeloid/erythroid ratios, and percentage of lymphocytes; the RBCs of the exposed mice 

also displayed increases in MCV, mean cell Hgb, and glycerol lysis time and in the incidence 

and severity of red cell morphological changes. The values for RBC, motor conduction velocity, 

mean corpuscular Hgb, and glycerol lysis time were statistically significant (p<0.05) for both 

males and females, whereas the remainder of the values were statistically significant for males 

only. 

Many of the effects occurred (and were statistically significant) by 14 days of exposure 

and most persisted throughout exposure, but the data provided indicated that the effects did not 

increase in severity with duration of exposure. Microscopic examination of the highconcentration group revealed changes in the thymus, bone marrow, lymph nodes, spleen, ovaries, 

59
and testes. The changes in the males occurred more often and with greater severity than those in 

the females. The most common compound-related histopathologic findings included myeloid 

hypoplasia of the bone marrow, lymphoid depletion of the periarteriolar sheath in the spleen and 

mesenteric lymph node, increased extramedullary hematopoiesis in the spleen, and plasma cell 

infiltration of the mandibular lymph node. The incidence and/or severity of thymic atrophy and 

myeloid hypoplasia appeared to increase with duration of exposure. The rats, not as severely 

affected as the mice, exhibited significantly (p<0.05) decreased WBC counts (males on day 14 

and females on days 14–91) and slightly decreased femoral marrow cellularity at 300 ppm (958 

mg/m3

). 

The investigators concluded that the hematologic effects observed in this study are 

related to concentration and are similar to those reported by other investigators, but that they are 

statistically significant only at 300 ppm (958 mg/m3

). In addition, the mouse is more sensitive 

than the rat to the effects of benzene, and male mice appear to be more sensitive than female 

mice. The experiment provided a LOAEL of 300 ppm (958 mg/m3

) and a NOAEL of 30 ppm 

(96 mg/m3

) for peripheral blood abnormalities in both rats and mice. The large numbers of 

animals used in the study and the evaluation of several parameters of hematotoxicity strengthen 

the conclusions of the investigators. 

4.2.1.4. Inhalation Exposure—Chronic Studies 

To further characterize the hematotoxicity of benzene, Snyder et al. (1978, 1980, 1982, 

1984, 1988) conducted a series of studies using either 100 or 300 ppm (319 or 958 mg/m3) 

benzene. These studies addressed mainly the influence of benzene concentration and animal 

species on toxic responses. The 300 ppm (958 mg/m3

) concentration was selected because it had 

produced severe hematotoxicity in AKR mice exposed for life, but this concentration had 

produced only lymphocytopenia in Sprague-Dawley rats exposed under the same conditions. 

Also of interest was whether benzene would affect the incidence of lymphoma in the different 

strains of rodent. The AKR strain carries a virus that produced a high incidence of spontaneous 

lymphoma, killing the mice within 1 year, and the C57BL strain carries a virus that yields a high 

incidence of lymphoma following exposure to radiation, carcinogens, and immunosuppressive 

agents (Snyder et al., 1980). 

Exposures were similar in all the studies. Essentially, test animals and control animals 

were placed in identical stainless steel inhalation chambers and exposed to benzene or filtered 

air, respectively, 6 hours/day, 5 days/week for life. Benzene concentrations, monitored every 30 

minutes by ultraviolet absorbance at 255 nm, were read from a calibration curve. Tail bleeding 

of test and control animals took place at least every 3–5 weeks during exposure and at about the 

same time of day for analysis of the following endpoints: WBC and RBC counts (Coulter), 

reticulocyte counts, differential counts, Hgb, and HCT. Evaluations of lactic acid dehydrogenase 

levels, RBC acetylcholinesterase levels, and reduced GSH levels were conducted less frequently. 

The evaluation also included cytogenetic analyses of bone marrow. Other tissues taken for 

microscopic examination included lung, liver, spleen, and kidney. 

In the first study of the series, Snyder et al. (1978) exposed AKR/J mice (60 test and 60 

control) and Sprague-Dawley rats (45 test and 27 control) to 300 ppm (958 mg/m3

) benzene. 

The rats had mild body weight depression that started at 30 weeks and persisted throughout life. 

60
The only significant hematologic findings in the rats (± 2 standard error [SE]) were 

lymphocytopenia and a trend to anemia. Lymphocyte counts for the exposed animals were 

approximately 75% of normal at 2 weeks and 53% of normal at 20 weeks. Because of increased 

mortality, blood analyses were discontinued after 1 year of exposure. The investigators also 

noted fatty changes in 77% of the bone marrow samples taken from the benzene-exposed 

animals compared with 42% of the controls. The last mouse in the study died after 28 weeks of 

exposure, probably of aplastic anemia (mean survival time, 11 weeks for test animals, 39 weeks 

for control animals). The treated animals suffered severe weight loss (59% weight change at 16 

weeks), marked lymphocytopenia (e.g., lymphocyte values that dropped to < 1000/mm3

 were 

~3% of control levels after 36 days of exposure), bone marrow hypoplasia (81% of exposed mice 

and 6% of controls, p<0.001 by chi-square), granulocytosis (fivefold increase after 9 weeks of 

exposure), and reticulocytosis. The blood analyses were suspended after 92 days of exposure 

due to early mortality. Neither species exhibited any signs of leukemia or preleukemia (the 

absence of overt signs of leukemia, characterized by bone marrow dysfunction, usually low 

numbers of certain cell types). The investigators suggested that this might be related to early 

mortality from aplastic anemia in the mice and a lack of opportunity for recovery of the bone 

marrow (a possible factor in leukemogenesis) in the rats. 

The investigators concluded from this study that inhalation of 300 ppm (958 mg/m3

) 

benzene causes significantly decreased survival, severe lymphocytopenia, and anemia 

accompanied by reticulocytosis and granulosis in AKR/J mice and lymphocytopenia, mild 

anemia, and moderately decreased survival in Sprague-Dawley rats. They further concluded that 

AKR/J mice are more susceptible to benzene-induced hematotoxicity than are Sprague-Dawley 

rats. 

The severity of hematologic effects induced by 300 ppm (958 mg/m3

) benzene in AKR/J 

mice prompted Snyder et al. (1980) to conduct a similar study with 50 AKR/J mice exposed to a 

lower concentration (100 ppm [319 mg/m3]) of benzene. The investigators also exposed 40 

C57BL/6 mice to 300 ppm (958 mg/m3) benzene to compare the susceptibility of a different 

strain of mice with that of rats. Hematologic parameters were measured in both strains at regular 

intervals during exposure. 

No statistically significant differences were seen between the AKR/J test and control 

mice in median survival (39 and 41 weeks, respectively), rate of weight gain, or in type and 

occurrence of lymphoma. Effects that did occur were not as severe as those observed at 300 

ppm (958 mg/m3

). But even at 100 ppm (319 mg/m3), the mice exhibited statistically 

significantly reduced lymphocyte counts after 1 week of exposure (cell counts were depressed to 

65% of control levels), and the depression persisted throughout exposure. RBC counts were also 

consistently depressed relative to controls, and the depression was statistically significant (± 2 

SE) at 9/19 monitoring periods. Neutrophils were elevated (± 2 SE) in 3/19 monitoring periods. 

Ten (20%) of the treated mice, but only one control, had bone marrow hypoplasia (p<0.05). 

C57BL/6 mice exposed to 300 ppm (958 mg/m3

) benzene had a median survival time of 

41 weeks, compared with 75 weeks for the controls. Starting 1 week after the beginning of 

exposure, mice exposed to benzene exhibited a statistically significant increase in incidence of 

lymphocytopenia at 29 of 30 monitoring periods (WBC counts were about 3000/mm3

 or ~15% of 

control values), and of anemia in 30 of 30 monitoring periods (RBC counts were ~80% of 

61
control values). After 17 weeks, the incidence of neutrophilia was significantly increased in 15 

of the 22 remaining monitoring periods. Throughout exposure, the peripheral RBCs exhibited 

anisocytosis and poikilocytosis. A neutrophilic shift to the left, characterized by the appearance 

of metamyelocytes, indicated immaturity. At the end of the first year, myelocytes, 

promyelocytes, and giant platelets began to appear. Blood analyses were discontinued after 61 

weeks because of decreased survival. Examination of the bone marrow revealed hyperplasia, 

mainly of granulopoietic elements, in 13 (33%) of the exposed mice; hyperplasia was not present 

in the controls. There was a significant increase in the incidence of hematopoietic neoplasms, 

including six cases (15%) of thymic lymphoma. Two control mice died with nonthymic 

lymphoma. 

The investigators concluded from this study that the two strains of mice are more 

sensitive than the Sprague-Dawley rat to the hematologic effects of benzene and that the AKR/J 

mouse is more sensitive than the C57BL/6 mouse. The response of the C57BL/6 mouse 

producing a proliferative effect on the myeloid cell line is significant, because this is the cell line 

that undergoes leukemic transformation in humans exposed to benzene. 

In the next study, Snyder et al. (1982) examined the effects of 300 ppm (958 mg/m3

) 

benzene on CD-1 mice, a strain not known to harbor any endogenous lymphoma virus. The 

effects were similar to those observed in the AKR/J and C57BL/6 strains and included decreased 

median survival time (25.5 weeks for benzene-exposed animals, 52.7 weeks for controls), 

statistically significantly depressed peripheral RBC and lymphocyte counts after the first week of 

exposure (± 2 SE), and statistically significant neutrophilia (± 2 SE) after 29 weeks (~209 days) 

of exposure and the appearance of Howell-Jolly bodies (remnants of nuclear chromatin seen in 

erythrocytes in certain anemic conditions) at 7 days, anisocytosis (erythrocytes showing 

excessive variation in size) at 22 days, and poikilocytosis (erythrocytes showing excessive 

variation in shape) at 92 days and a shift to immature myeloid cells at 217 days. 

Tumor incidence was not significant. One benzene-exposed animal developed acute 

myeloblastic leukemia, one had chronic myelogenous leukemia, one had a benign lung adenoma, 

and two had malignant lymphoma with thymic involvement. Of the 35 benzene-exposed animals 

that did not have neoplasia, 9 had bone marrow hyperplasia, 11 had bone marrow hypoplasia, 6 

had splenic hemosiderin pigments (indicating hemolysis or ineffective erythropoiesis), and 19 

had splenic hyperplasia. The investigators noted that “exposed mice dying with bone marrow 

hypoplasia survived, on the average, 75 fewer days than mice dying with bone marrow 

hyperplasia. This may indicate that bone marrow hypoplasia is an early response to exposure 

followed by bone marrow hyperplasia. It may also indicate two different responses to the 

exposure.” 

Snyder et al. (1988) examined the influence of modifications in inhalation protocol on 

benzene-induced hematotoxicity and tumor development in C57BL/6 and CD-1 mice. Exposure 

was by two protocols, one representing intermittent occupational exposures (1 week of exposure 

to 300 ppm [958 mg/m3

] benzene followed by 2 weeks of nonexposure alternately for life), and 

the other representing intense but short-term exposures (1200 ppm [3834 mg/m3] benzene for 10 

weeks); all exposures were 6 hours/day, 5 days/week. The long-term intermittent exposures 

produced earlier mortality in both strains of mice than did the short-term, 10-week exposures. 

Both types of exposure produced severe lymphocytopenia and moderate anemia in both strains 

62
of mice, but peripheral blood values in the 1200 ppm (3834 mg/m3

)/10-week exposure groups 

returned to normal after termination of exposure. The 300 ppm (958 mg/m3) exposure induced 

lymphocytopenia and anemia throughout the course of exposures. Tumor incidences were 

significantly increased (p<0.05–0.001) with both protocols in CD-1 mice and with the 

intermittent, long-term exposure in the C57BL/6 strain. A puzzling finding was that neither 

leukemia nor lymphoma occurred in either strain by either protocol. The investigators concluded 

that benzene was cytotoxic by either protocol, especially to circulating lymphocytes (but 

recovery took place when exposure to the high concentration ceased) and that both protocols did 

induce tumors. The Snyder et al. (1978, 1980, 1982, 1984, 1988) studies were designed to 

answer specific questions and used only one concentration at a time (except for the study on 

exposure protocol), and therefore they were not sufficient for the derivation of LOAEL or 

NOAEL. 

4.2.1.5. Effects on Stem Cell Populations 

The benzene-induced peripheral blood abnormalities reflect a disruption of 

hematopoiesis in the bone marrow; thus many experimental studies have focused on this process, 

analyzing the total cellularity (the total number of bone marrow cells obtained from a femur), as 

well as the individual cellular components of the bone marrow. The pluripotent hematopoietic 

stem cells are capable of self-renewal and can differentiate along at least three lineages: 

erythrocytic, granulocytic/macrophagic, and thrombocytic (Snyder, 1987). These cells are 

quantitated by the spleen colony assay in which the colonies formed by the cells in the spleens of 

lethally irradiated host animals are counted; these are called colony-forming units–spleen (CFUS). The specific progenitor cells, in contrast to the pluripotential stem cells, are committed to 

differentiate, and they also have some capacity for self-renewal (cell division). These cells are 

enumerated by counting the colonies formed in cell cultures grown in the presence of specific 

growth-factor stimuli (Snyder, 1987). The two types of erythroid progenitor cells are the lessdifferentiated burst-forming units–erythroid (BFU-E), which responds to a factor called “burstpromoting activity” or to IL-2, and the more differentiated erythroid progenitor, the colonyforming units–erythroid (CFU-E), which responds to erythropoietin. The 

granulocyte/macrophage colony-forming units–culture (GM-CFU-C) respond to conditioned 

media from human leukemia cells or from organs of animals treated with plant lectins. The Tand B-lymphocytes arise from the pluripotential hematopoietic stem cells in the bone marrow. 

The cells progress to the recognizable blast cells and their progeny and further mature to the 

functional forms normally observed in peripheral blood (Snyder, 1987). 

Seidel et al. (1989) examined the effects of benzene on peripheral blood cell counts and 

the hematopoietic stem cell compartments of the bone marrow. Female BDF1 mice were 

exposed to 100, 300, or 900 ppm (319, 958, or 2875 mg/m3

) benzene 6 hours/day, 5 days/week 

for 16 weeks. Benzene concentrations were monitored twice a day. Control mice, placed in 

chambers, were exposed to air. The evaluation of hematopoietic effects included WBC 

(Coulter), reticulocyte, and differential counts and assays for CF-U-S, BF-U-E, CFU-E, and 

GM/(CFU-C) using standard procedures. The assays, conducted about every 2 weeks, usually 

took place about 6 hours after the end of the exposure period of the preceding week. 

There was a dose-dependent reduction in lymphocyte count after both 4 and 8 weeks of 

exposure. Lymphocyte counts of the animals exposed to 300 and 900 ppm (958 and 2875 

63
mg/m3

) returned to control values 1–2 weeks after cessation of exposure, whereas the 

lymphocytes in the 100 ppm (319 mg/m3) group remained low during that time. The animals 

exposed to 300 and 900 ppm (958 and 2875 mg/m3) developed a slight anemia that did not 

worsen with longer exposure. Three-fourths of the animals with elevated CFU-E in the bone 

marrow also showed reticulocytosis at 13 weeks. The number of granulocytes was apparently 

not affected. The investigators noted that control values for the 300 ppm (958 mg/m3

) groups 

were approximately twice those for the 100 and 900 ppm (319 and 2875 mg/m3) groups, possibly 

because the 300 ppm (958 mg/m3

) experiment had been performed 1 year earlier. 

Assays performed after 4 and 8 weeks of exposure demonstrated decreases in stem cell 

numbers that were dose related, particularly at 8 weeks. At 4 weeks, the CFU-E count was 

depressed more than other stem cell counts. At 100 ppm (319 mg/m3

), the CFU-E were 

depressed to 48% of control, whereas the BFU-E, CFU-S, and CFU-C (colony-forming 

unit–culture) were depressed to 88%, 92%, and 94% of control, respectively. At 8 weeks, the 

values for CFU-E in animals exposed to 100, 300, and 900 ppm (319, 958, and 2875 mg/m3

) 

were 99%, 73%, and 35% of controls, respectively. Similar patterns of depression were 

observed for BFU-E and CFU-C numbers of animals exposed to benzene for 8 weeks, and CFUS numbers were depressed to 65% of control at 300 ppm (958 mg/m3

), the only dose tested after 

8 weeks of exposure. Recovery studies demonstrated a gradual regeneration of progenitor cell 

numbers, and animals exposed to 300 ppm (958 mg/m3) for 16 weeks had completely recovered 

by 73–185 days postexposure. The investigators concluded that the erythroid system is more 

sensitive than the myeloid cell system to the effects of benzene, with the CFU-E showing more 

sensitivity than other components of the erythroid system under the conditions of this study. 

Two studies were designed to examine the effects of low concentrations of benzene on 

the erythroid and myeloid progenitor cells (Baarson et al., 1984; Baarson and Snyder, 1991). 

Baarson et al. (1984) examined the effects of subchronic exposure to a low concentration of 

benzene on the erythroid progenitor cells in the bone marrow and spleen of mice. Male 

C57BL/6J mice inhaled 10 ppm (32 mg/m3

) benzene vapor 6 hours/day, 5 days/week for as long 

as 178 days. Peripheral blood counts and bone marrow and spleen BFU-E and CFU-E assays 

were performed on days 32, 66, and 178 of exposure. The numbers of circulating RBCs in the 

benzene-treated groups were significantly decreased (p<0.05) at both 66 and 178 days. 

Lymphocyte values were decreased (p<0.05) at all three time points. Benzene did not affect 

levels of circulating neutrophils, bone marrow cellularity, and numbers of nucleated red cells in 

the bone marrow. 

The numbers of bone marrow CFU-E gradually declined during exposure to only 5% of 

control values after 178 days. The decline, statistically significant (p<0.01) at all three time 

points, was apparently exponential with a half-life of about 2 months. Bone marrow BFU-E 

colonies were significantly depressed in benzene-exposed animals (to about 55% of control; 

p<0.01) at 66 days, but recovered to control values by 178 days. Depressions in splenic 

nucleated red cell numbers (to about 15% of control values; p<0.05) and in splenic nucleated 

cellularity (p<0.05) occurred at day 178. Splenic CFU-E colonies in benzene-exposed mice 

were decreased to 10% of control values at 178 days (p<0.05). In contrast, splenic BFU-E 

colonies were increased (but not significantly) at all time points. The spleen is a site of 

extramedullary erythropoiesis in the rodent, particularly under stressful conditions. The 

reduction in the numbers of nucleated red cells and CFU-E in the spleen following exposure to 

benzene indicated that the compensatory mechanism of the spleen was also affected. The 

64
investigators concluded that low concentrations of benzene are hematotoxic and that further 

studies to examine this effect would be warranted. 

In the later study, Baarson and Snyder (1991) reported that the ingestion of ethanol in 

combination with exposure to benzene may enhance benzene-induced toxicity of the erythroid 

progenitor cells in C57BL/6J mice. These studies identify a LOAEL of 10 ppm (32 mg/m3) for 

depressed hematopoiesis in mice. 

Vacha and coworkers (1990) exposed female C57BL/6×DBA/2F1 hybrid mice to 0 or 

300 ppm (0 or 958 mg/m3

) benzene 6 hours/day, 5 days/week for 6–7 weeks. Indices of 

hematopoiesis were measured in peripheral blood (RBC and WBC count, Hgb, HCT, 

reticulocyte, and leukocyte count) in addition to 59Fe accumulation in the erythropoiesis organs 

(spleen and bone marrow) and in the peripheral RBCs. The distribution of developmental 

classes of erythroblasts was also determined. This study found that animals became anemic after 

6–7 weeks of benzene exposure. The number of erythroblasts in the bone marrow was not 

different; however, exposure to benzene shifted the population to a less mature class of cells. 

The number of colonies derived from BFU-E and CFU-E were decreased to 70 and 34% of 

controls, respectively. A LOAEL of 300 ppm (958 mg/m3

) was established for hematotoxic 

effects. 

Toft et al. (1982) evaluated the adverse effects of occupationally relevant levels of 

benzene on the bone marrow of mice. Male NMRI mice inhaled concentrations of benzene that 

ranged from 1 to 200 ppm (3.2 to 639 mg/m3

). Exposures were either continuous (24 hours/day 

for 4 to 10 days) or intermittent (8 hours/day, 5 days/week for 2 weeks) using five mice per 

group. Endpoints included the number of nucleated cells per tibia, the number of colonyforming granulopoietic stem cells (CFU-C/tibia), and frequency of micronuclei in polychromatic 

erythrocytes (PCE). Mice exposed continuously to 21 ppm (67 mg/m3

) benzene exhibited 

significant and concentration-dependent alterations in all three parameters. The values after 4 

days of exposure to 21 ppm (67 mg/m3

) of benzene (estimated from a graph) were cells/tibia, 

~24% of control; CFU-C/tibia, ~32% of control; micronuclei/500 PCE, 6 (control value, 

0.41/500). 

Furthermore, intermittent inhalation exposure to 1.0, 10.5, 21, 50, 95, and 107 ppm (3.2, 

33, 67, 160, 303, and 342 mg/m3

) for 2 weeks produced concentration-dependent decreases in 

CFU-C/tibia and increases in the frequency of micronuclei, significant at 21 ppm (67 mg/m3

), 

and concentration-dependent increases in cellularity, significant at 50 ppm (160 mg/m3). 

Animals exposed intermittently to 95 or 201 ppm (303 or 642 mg/m3) benzene 2–8 hours/day, 5 

days/week for 2 weeks had significant decreases in cellularity and CFU-C/tibia (95 ppm [303 

mg/m3

] for 6 and 8 hours/day) and increases in the frequency of micronuclei (95 ppm [303 

mg/m3] for 4 hours). At 201 ppm (642 mg/m3), all exposure protocols, except for the 2 

hours/day exposure, produced significantly adverse effects on all parameters. Statistical 

significance was determined by the Student’s t test for cellularity and CFU-C and by the WilcoxWhitney method for micronuclei, which was set at p0.05. Statistical values were not given for 

the individual data points. 

The investigators concluded that the CFU-C/tibia was suppressed to a greater extent than 

was the overall cellularity for most exposures and that CFU-C are more sensitive to prolonged 

65
exposure to low levels of benzene than are the majority of bone marrow cells. However, higher 

exposures for short durations produced the reverse, suppressing cellularity but not CFU-C. This 

could indicate that short exposures cause rapid injury to the bone marrow but that the injury 

could be offset by a compensatory increase in the proliferation rate of the bone marrow. The 

investigators further concluded that intermittent exposure is less effective than continuous 

exposure, that the induction of micronuclei is important because somatic mutations precede most 

chemically induced cancers, and that the cytotoxicity observed in the study could lead to aplastic 

anemia. 

Because of the relationship between exposure to benzene and the development of 

myeloblastic leukemia, Dempster and Snyder (1990) evaluated the possibility that exposure 

provides a growth advantage for granulocytic progenitor cells over erythroid progenitor cells. 

Basically, DBA/2 male mice inhaled benzene concentrations of 0, 10, 30, or 100 ppm (0, 32, 96, 

or 319 mg/m3

) 6 hours/day, 5 days/week for 5 days and were evaluated for BFU-E, CFU-E, and 

GM-CFU-C on days 1 and 5 after exposure. To determine the effects of benzene on stem cells 

during increased erythropoiesis, subgroups of five exposed mice received injections of hemolytic 

doses of phenylhydrazine during or after exposure. Controls were exposed to air and/or injected 

with saline. Brief descriptions of treatment and the effects, significant at p<0.05, are as follows. 

Benzene only. One day after exposure, there was a dose-dependent depression in bone 

marrow BFU-E and CFU-E, but GM-CFU-C were not affected (except for a spurious increase at 

30 ppm [96 mg/m3

]). Splenic CFU-E and BFU-E were increased at 10 ppm (32 mg/m3) and 

CFU-E were depressed at 100 ppm (319 mg/m3

). Splenic granulocytic cells were unaffected. 

Five days after exposure, BFU-E and CFU-F were the same as in controls, but GM-CFU-C were 

decreased at 10 and 100 ppm (32 and 319 mg/m3); splenic CFU-E exhibited a concentrationrelated increase, significant at 30 and 100 ppm (96 and 319 mg/m3

), whereas the numbers of the 

other two progenitor cell types were normal. 

Benzene + phenylhydrazine. One day after exposure, bone marrow BFU-E were 

depressed at 30 and 100 ppm (96 and 319 mg/m3

), and CFU-E were depressed only at 30 ppm 

(96 mg/m3); there was no effect on the number of GM-CFU-C. Splenic BFU-E and GM-CFU-C 

showed concentration-dependent depressions, and CFU-E were depressed, but only at 100 ppm 

(319 mg/m3). Five days after exposure, bone marrow CFU-E were elevated in mice exposed to 

10 ppm (32 mg/m3

), and GM-CFU-C exhibited a concentration-dependent depression; bone 

marrow BFU-E counts were similar to control counts. Splenic GM-CFU-C were elevated at all 

concentrations, CFU-E were elevated at 10 and 30 ppm (32 and 96 mg/m3

) and depressed at 100 

ppm (319 mg/m3), and BFU-E were elevated only at 10 ppm (32 mg/m3). 

The investigators concluded that acute exposure to benzene has different effects on 

erythroid and granulocytic progenitor cell populations, which resulted in a growth advantage for 

granulocytic cells in both the bone marrow and the spleen of exposed mice. This observed shift 

toward granulopoiesis occurred even in mice treated with the erythropoiesis stimulus 

phenylhydrazine, but the effects were short-lived. The bone marrow erythroid progenitor cells 

had recovered from their depression 5 days after exposure, and at that time the granulocytic 

progenitor cells had become depressed. The dose-dependent increase in splenic CFU-E 5 days 

after exposure probably reflects the spleen’s attempt to repopulate the erythron. 

66
In a series of studies, Luke et al. (1988a, b) and Tice et al. (1989) examined the effect of 

sex, route, schedule, and duration of exposure on benzene-induced cytotoxicity and genotoxicity 

of the bone marrow in mice. The effects on percentage of PCE in peripheral blood, PCV, and 

bone marrow cellularity (all considered by the investigators to be measures of bone marrow 

cytotoxicity) are reviewed below. 

Groups of 6–10 male and female DBA/2 mice, male B6C3F1 mice, and male C57BL/6J 

mice inhaled either 300 ppm (958 mg/m3

) benzene or ambient air 6 hours/day for 13 weeks. 

Exposures took place on either 3 consecutive days/week or 5 consecutive days/week. For 

comparison with the oral route of exposure, one group of B6C3F1 males received doses of 400 

mg/kg by gavage on 5 consecutive days/week for 14 weeks. The oral dose of 400 mg/kg was 

estimated to exceed the total amount of benzene absorbed by a mouse during a 6-hour exposure 

to 300 ppm (958 mg/m3

). Peripheral blood smears, prepared weekly, were used to examine PCE. 

PCVs and bone marrow cellularity were determined at the end of the study. The data for the 

three parameters were statistically analyzed using temporal averages and then evaluated for 

differences related to sex, strain, regimen, and route of exposure using a two-way Brown 

Forsythe analysis of variance (ANOVA). Group mean data were compared using Student’s t 

test. 

Benzene initially induced a significant depression in the number of PCE in the peripheral 

blood of all three strains exposed by either route and both inhalation regimens. However, the 

extent and duration of the depression varied by sex, strain, and exposure regimen. Female 

DBA/2 mice exhibited the smallest initial decrease in PCE, and values returned to control levels 

by week 3 of exposure. The extent of depression was not dependent on exposure regimen. 

Among the males of the different strains, the DBA/2 mouse was the most severely affected, 

particularly when exposed to benzene 3 days/week. ANOVA revealed significant differences in 

the ability of benzene to suppress erythropoiesis that were related to both strain (for C57BL/6 

and DBA/2 mice) and exposure regimen. Oral treatment produced an initial suppression in PCE 

levels in the male B6C3F1 mice that persisted for only 2 weeks and was not revealed in the 

temporal average. 

The study also demonstrated a significant depression in the PCV and bone marrow 

cellularity (determined at the end of the study) of the exposed mice. With regard to the PCV 

values, the depression was independent of the inhalation regimen in B6C3F1 and C57BL/6 mice. 

After correcting for differences in PCV levels in controls, the data showed that inhalation 

produced a greater effect than oral exposure to benzene, the extent of which was straindependent for all three strains (B6C3F1  C57BL/6 > DBA/2; p<0.0163). In DBA/2 mice the 

extent of depression was sex dependent (male > female; p<0.025). With regard to bone marrow 

cellularity, the depression observed after inhalation exposure depended on sex (male > female; 

p<0.0002) and strain (DBA/2 < B6C3F1  C57BL/6; p<0.0012). Among the male mice, the 

extent of depression was dependent on route (inhalation > oral; p<0.012) and was not dependent 

on regimen. 

These studies used rather small numbers of animals but otherwise appear to have been 

thoroughly conducted and evaluated; however, the information provided is for exposure to 

benzene that is significantly above the most recent occupational limits. The investigators 

67
suggest that further studies are needed to evaluate the influences on benzene toxicity relative to 

regimen and route of exposure at lower doses. 

In a series of experiments, Cronkite et al. (1985, 1989) examined the hematotoxicity of 

inhaled benzene in mice. The study examined the effects on and recovery of peripheral blood, 

bone marrow, and progenitor cells, as well as the development of neoplasia and the influence of 

exposure regimen on toxicity. In the first study, male and female C57BL/6 BNL mice were 

exposed for 6 hours/day, 5 days/week to benzene concentrations of 10, 25, 100, or 400 ppm (32, 

80, 319, or 1278 mg/m3

) for 2 weeks or to 300 ppm (958 mg/m3) for durations of exposure 

ranging from 2 to 16 weeks. The endpoints of toxicity included WBC, RBC (Coulter), and 

differential counts; HCT; bone marrow cellularity; CFU-S in the bone marrow and spleen; and 

tumor development. Assays were performed 18–22 hours after termination of exposure or after 

various periods of recovery up to 16 weeks. 

Inhalation of 25, 100, or 400 ppm (80, 319, or 1278 mg/m3

) benzene for 2 weeks did not 

affect granulocyte counts but did result in a concentration-related depression in the number of 

lymphocytes in the peripheral blood; there was no effect on lymphocytes at 10 ppm (32 mg/m3

). 

HCT values were decreased at 100 and 400 ppm (319 and 1278 mg/m3); RBC counts were also 

decreased, but no data were given. At the end of exposure, lymphocyte depression ranged from 

about 20 to 65% of control for the three highest exposure groups. For mice exposed to 300 ppm 

for 2 weeks, lymphocyte levels were similar to control levels by 4 weeks after the end of 

exposure, and recovery of the remaining exposure duration groups was complete by 8 weeks 

after exposure. This experiment provided an estimated LOAEL of 25 ppm (80 mg/m3) and a 

NOAEL of 10 ppm (32 mg/m3

) for lymphocyte depression in mice exposed to benzene for 2 

weeks (Cronkite et al., 1985). These results differed from those of Ward et al. (1985), which 

suggested a LOAEL of 300 ppm (958 mg/m3

) for all hematologic parameters tested in both mice 

and rats. The main difference in the two studies was the use of different strains of mouse. 

Cronkite et al. (1985) also determined bone marrow cellularity, the percentage of bone 

marrow stem cells synthesizing DNA, numbers of pluripotent stem cells in bone marrow, and 

CFU-S. Cellularity was determined in animals inhaling 0, 10, 25, or 100 ppm (0, 32, 128, or 319 

mg/m3

) for 2 weeks (the 400 ppm [1278 mg/m3] concentration was not tested as it was for the 

other parameters of the study). Benzene levels of 100 ppm (319 mg/m3) significantly depressed 

bone marrow cellularity and stem cell CFU-S numbers (p<0.003 and p<0.001, respectively) in 

exposed mice. In contrast, 10 and 25 ppm (32 and 128 mg/m3) had no effect on these 

parameters. A single assay demonstrated a depression in the fraction of stem cells in DNA 

synthesis in mice exposed to 10 ppm (32 mg/m3). There was no effect at 25 ppm (128 mg/m3), 

and slight increases in the fraction in DNA synthesis at 100 and 400 ppm (319 and 1278 mg/m3

) 

(p<0.17 and p<0.08, respectively). These results for DNA synthesis are of uncertain relevance. 

In another experiment, the recovery of bone marrow CFU-S population was tracked after 

exposure of the animals to 300 ppm (958 mg/m3

) for 2, 4, 8, or 16 weeks. In animals exposed 

for 2 and 4 weeks, bone marrow CFU-S numbers were depressed to 90% of control values at the 

end of exposure but had recovered to >100 and >95% of control values, respectively, 4 weeks 

later. In animals exposed for 8 weeks, CFU-S levels were decreased to 50% of control values 4 

weeks after the last exposure, but they recovered 8 weeks later (after 16 weeks, the values were 

>100% of control). In animals exposed for 16 weeks, CFU-S values were depressed to 27% of 

68
control, and although recovery was occurring by week 2 after the last exposure, it was still not 

complete by 16 weeks after termination of exposure. The investigators concluded that exposure 

to 300 ppm (958 mg/m3

) benzene for 16 weeks significantly depressed the hemopoietic stem 

cells and that recovery was incomplete 16 weeks after termination of exposure. 

In reviewing the effects of benzene on the CFU-S compartment, Snyder (1987) pointed 

out that under normal circumstances the CFU-S cells are nondividing and in that state are 

probably resistant to benzene toxicity; however, in response to benzene-induced injury of more 

differentiated cell types, the CFU-S may be drawn into the cell cycle and thereby become 

susceptible to benzene. 

Cronkite et al. (1985) also noted that animals exposed to 300 ppm (958 mg/m3

) benzene 

for 16 weeks began to die from thymic and nonthymic lymphomata and solid tumors, beginning 

at 330 days of age. In a continuation of this study, Cronkite et al. (1989) (1) evaluated the 

effects of benzene on erythrocyte progenitors (BFU-E, CFU-E) and on neutrophil and 

macrophage progenitors (GM-CFU-C), (2) compared the effects of two different exposure 

regimens on peripheral blood cell counts and progenitor cell numbers, (3) evaluated mortality 

and neoplasia, and (4) tested the functional capacity of the stem cells using a “rescue assay.” 

Hale-Stoner mice, aged 12–14 weeks, were used in the progenitor cell assays. Bone marrow cell 

suspensions harvested from mice exposed to 400 ppm (1278 mg/m3

) benzene 6 hours/day, 5 

days/week for 9.5 weeks and from their corresponding controls were counted and incubated to 

form plasma clots. The cells used in the CFU-E assay were cultured in 2% fetal calf serum; the 

cells used in the BFU-E assay were cultured in erythropoietin and pokeweed-mitogen 

conditioned medium. The clots were scored for erythrocytic colonies (CFU-E) or 

megakaryocytes (BFU-E). The cells used in the CFU-G and GM-CFU assays were cultured in 

agar for subsequent scoring. Progenitor cell assays resulted in the following: CFU-E numbers 

did not change significantly after 4 days of exposures, but the numbers had decreased 

significantly (to approximately 45–70% of control values) by assay days 29, 48, and 65. Twelve 

days after exposure ended, the cell numbers had recovered to 200% of control values. The BFUE values also decreased significantly (to just a few percentage points of control values) at 29, 48, 

and 65 days, but they had recovered to only about 40% of control values on day 12 after 

exposure. During the time of this assay, the peripheral RBC counts were decreased to about 

50–75% of control values. 

Granulocyte/macrophage aggregates, also observed in the clots, were decreased to about 

10–15% of control values by day 29 and had recovered to about 85% of control values by 12 

days after exposure. The GM-CFU cells in an agar assay decreased to approximately 40% of 

control values after 30 days of exposure, increased to about 75% by termination of exposure at 

day 65, decreased to approximately 60% by 72 days, and then recovered to almost 80% by 80 

days. The numbers of granulocytes in the blood also fluctuated during this time. 

Influence of exposure protocols. CBA/Ca BNL male mice were exposed to either 316 

ppm (1010 mg/m3

) benzene administered 6 hours/day, 5 days/week for 19 exposures or to 3000 

ppm (9584 mg/m3) benzene administered 6 hours/day for 2 successive days (Cronkite et al., 

1989). After 19 exposures, 316 ppm (1010 mg/m3

) benzene had reduced lymphocyte counts 

from 7500/L to 300/L, and 3000 ppm (9584 mg/m3

) for 2 days had resulted in a reduction 

from 6700/L to 3300/L 1 day after the end of the exposure. The differences in the effects of 

69
the two regimens still persisted at 214 days after exposure. Neutrophil counts of the group 

exposed to 316 ppm (1010 mg/m3

) were also depressed more than those exposed to 3000 ppm 

(9584 mg/m3). By 214 days after exposure, the neutrophil counts in the group treated with 3000 

ppm (9584 mg/m3

) had recovered, whereas those in the group treated with 300 ppm (958 mg/m3) 

had not. All leukocyte cell types were decreased in number as compared with those in shamexposed mice on day 1, and all absolute counts except for the large unstained cells (mostly large 

lymphocytes) were depressed on 32, 67, and 214 days after the last exposure. These data were 

not given for the animals exposed to 3000 ppm (9584 mg/m3

). 

On day 1 after exposure, both treatments had induced a reduction in bone marrow 

cellularity; this was significantly greater in the animals exposed to 316 ppm (1010 mg/m3

) for 19 

exposures (depression to ~33% of control value) than in those exposed to 3000 ppm (9584 

mg/m3

) for 2 days (depression to ~59% of control value). Cellularity had returned to control 

values by 32, 67, and 214 days after exposure. The CFU-S values followed the same response 

pattern except that by day 214 after exposure recovery was incomplete for the group exposed to 

316 ppm (958 mg/m3

). 

In another experiment, exposure to 300 ppm (958 mg/m3

) benzene for 80 weeks produced 

a severe depression in CFU-S (to 38% of control) that had not completely recovered 178 days 

after exposure ended. Exposure to 3000 ppm (9584 mg/m3) for 8 weeks produced a milder 

depression in CFU-S (to 62% of control) that had recovered by 30 days. Decreases in bone 

marrow cellularity were nearly the same for the two exposure scenarios. 

Rescue assay. The injection of normal bone marrow cells into lethally irradiated rodents 

enhances their survival (Cronkite et al., 1989). The serial rescue assay tested the effect of 

benzene on this particular function of the bone marrow. C57BL/6BNL male and female mice 

were irradiated with 850 rad, then injected with bone marrow cells from mice that had been 

exposed to 300 ppm (958 mg/m3

) benzene 6 hours/day, 5 days/week for 16 weeks 752 days 

earlier. Thirty days after the initial transplants, some of the survivors’ marrow was injected into 

another group of irradiated mice for a secondary transplant, and 30 days later those recipients 

served as donors in the third transplant to still another group of irradiated mice. The recipients 

of bone marrow in each transplantation of the series were observed for up to 594 days. In the 

first rescue, the 30-day survival was 100%. At 290 days, survival of the mice that received cells 

from control mice was 93%, and survival of the mice that received cells from benzene-exposed 

mice was 64%. 

In the secondary transplant, no mice died by 30 days, but the mice that received control 

cells started dying at day 171 and mice that received cells from donors exposed to benzene 

started to die at day 70. In the tertiary rescue, 50% of the recipients of bone marrow from 

benzene-exposed donors were dead by day 30 and 95% were dead by day 245. Among the 

animals that received cells from the sham-exposed mice, only one died by day 30 and 50% died 

by day 375. The last animal died on day 594. The mechanism for the stem cell malfunction that 

led to increased mortality is not clear, but the investigators suggest that the observation could be 

associated either with genetic injury to the stem cells, exhaustion of the G0 hematopoietic stem 

cells, or a radiation effect on the stroma of the recipient. 

70
BDF1 mice were exposed to 0, 100, 300, or 900 ppm (0, 319, 958, or 2875 mg/m3

) 

benzene for up to 4 weeks (Seidel et al., 1990). The numbers of hematopoietic progenitor cells, 

early and late progenitors (BFU-E, CFU-E), and granuloid progenitors (CFU-C) were 

determined. A group was generated to establish the effect of ethanol (drinking water) on these 

effects. This study demonstrated that benzene decreased the number of CFU-E per femur in a 

concentration-dependent manner. This effect was evident at 300 and 900 ppm (958 and 2875 

mg/m3

) concentration; however, the effect in the 100 ppm (319 mg/m3) exposure group was 

uncertain, as the study focused on the effects of ethanol on benzene toxicity. The 

LOAEL/NOAEL was thus difficult to determine. 

Male Sprague-Dawley rats (40/group) were exposed to vapor concentrations of 0 or 100 

ppm (0 or 319 mg/m3

) benzene 6 hours/day, 5 days/week for life (American Petroleum Institute, 

1983). Blood samples were obtained at 2- to 4-week intervals throughout the treatment period. 

The erythrocyte and lymphocyte counts were depressed at nearly every sampling time in treated 

rats, but the extent of decrease was rarely statistically significant. Nonetheless, the frequencies 

of depression were highly significant for each cell type (p<0.002, Wilcoxon Rank Sum Test), 

and the overall results were interpreted as evidence of hematotoxicity. Significantly increased 

incidences of splenic hyperplasia (p<0.005) and hemosiderin pigments (p<0.001) were observed 

in benzene-exposed rats. The incidences of normally rare tumors in treated rats were 4/40 in 

liver, 2/40 in Zymbal gland, and 1/40 chronic myelogenous leukemias. The authors considered 

these tumors to be related to the benzene exposure. This study identifies a LOAEL of 100 ppm 

(319 mg/m3

) for hematologic effects in rats. 

Decreases in bone marrow cellularity in the femur, HCT, and leukocytes were seen in 

DBA/2 mice (20/group) exposed to 300 ppm (958 mg/m3

) benzene in air 6 hours/day, 5 

days/week for 2 weeks (Chertkov et al., 1992). In most cases no erythroid or myeloid 

clonogenic cells could be recovered in bone marrow cultures started after the last benzene 

exposure. After 2 weeks of recovery, however, body weight, HCT, bone marrow cellularity, and 

committed hematopoietic progenitor cells had recovered to near normal values. 

Neun et al. (1992, 1994) compared the in vivo sensitivity of Swiss-Webster and 

C57B1/6J mice to inhaled benzene with the in vitro sensitivity of culture bone marrow cells to 

benzene metabolites. Mice were exposed to 300 ppm (958 mg/m3

) benzene 6 hours/day, 4 

days/week for 2 weeks. Swiss-Webster mice were more sensitive to benzene exposure, as 

indicated by much greater reductions in femoral bone marrow cellularity and in the number of 

CFU-E per femur after in vivo exposure. Neither phenol nor muconic acid were toxic to cultured 

CFU-E from either mouse strain. CFU-E from Swiss-Webster mice were more sensitive than 

CFU-E from C57B1/6J mice to 1,4-benzoquinone or hydroquinone. Thus, both in vitro and in 

vivo data indicated that Swiss-Webster mice were more sensitive to benzene toxicity than were 

C57B1/6J mice. 

Female BDF1 mice (C57BL/6 × DBA/2F1 hybrids) were exposed to 0, 100, 300, or 900 

ppm (0, 319, 958, or 2875 mg/m3

) benzene 6 hours/day, 5 days/week for up to 8 weeks (Plappert 

et al., 1994). Hematologic studies included peripheral blood data, T4 and T8 lymphocyte counts 

in the blood and spleen, and hematopoietic stem and progenitor cell assays in the marrow CFUS, CFU-C, and BFU-E, CFU-E. No significant changes were observed in the peripheral blood 

data of mice exposed to 900 ppm (2875 mg/m3) benzene. Some perturbation of the reticulocyte 

71
numbers was observed, but values at days 3 and 5 did not differ from controls. Absolute 

numbers of lymphocytes and neutrophils did not differ from controls (no data shown). Slight 

anemia was observed at 4 and 8 weeks of treatment with 300 and 900 ppm (958 and 2875 

mg/m3) benzene. Minor changes occurred in the stem and progenitor cells. CFU-E depression 

after 4 days of exposure was significant. A dose-dependent depression of colony-forming cell 

number appeared at 4 and 8 weeks of exposure, with maximal effect at the level of CFU-E. 

Farris et al. (1997a) exposed male B6C3F1 mice to 0, 1, 10, 100, or 200 ppm (0, 3.2, 32, 

319, or 639 mg/m3

) benzene 6 hours/day, 5 days/week for 1, 2, 4, or 8 weeks. A separate 4-week 

experiment at benzene concentrations of 0, 1, 5, and 10 ppm (0, 3.2, 16, and 32 mg/m3

) was also 

conducted. Another group of animals was exposed to benzene for 4 weeks and then allowed to 

recover for up to 25 days. Bone marrow cell counts, high proliferative potential colony-forming 

unit (CFU-HPP), percentage of CFU-HPP cells in S-phase, bone marrow CFU-E, bone marrow 

GM-CFU, labeling index, B and T lymphocytes, and hematology parameters were determined. 

Data sets for each parameter were individually evaluated with a univariate two-way ANOVA 

using contrasts to determine treatment effect. Statistical significance was determined at a level 

of p<0.05. 

There were no significant effects on hematopoietic parameters from exposure to 10 ppm 

(32 mg/m3

) benzene or less. Exposure of mice to 100 or 200 ppm (320 or 640 mg/m3) reduced 

the number of total bone marrow cells, progenitor cells, differentiating hematopoietic cells, and 

most blood parameters, with a concentration-related effect. Replication of primitive progenitor 

cells in the bone marrow was increased during the exposure period, possibly in compensation for 

benzene cytotoxicity. Recovery after benzene exposure was rapid. Most hematopoietic 

parameters returned to control levels within 4 days following the 100 ppm (320 mg/m3) exposure 

and within 11 days following the 200 ppm (640 mg/m3) exposure. The percentage of CFU-HPP 

in S-phase was elevated throughout the 25-day recovery period evaluated. Farris et al. (1996) 

reported that the frequency of micronucleated erythrocytes was increased in B6C3F1 male mice 

exposed to 100 and 200 ppm (320 and 640 mg/m3) benzene for 8 weeks. The authors suggested 

that the increased proliferation of primitive progenitor cells, in concert with genetic damage, 

provides the components for producing an increased incidence of lymphoma in mice. The 

effects on hematopoietic parameters indicate a LOAEL of 100 ppm (320 mg/m3

) and a NOAEL 

of 10 ppm (32 mg/m3). 

Farris et al. (1997b) exposed male B6C3F1 mice to 0, 1, 10, 100, or 200 ppm (0, 3.2, 32, 

319, or 639 mg/m3) benzene 6 hours/day, 5 days/week for 1, 2, 4, or 8 weeks. A separate 4-week 

experiment at benzene concentrations of 0, 1, 5, and 10 ppm (0, 3.2, 16, and 32 mg/m3) was also 

conducted. Spleen and thymic lymphocyte counts were determined for three mice in each 

treatment. Femoral B lymphocyte counts were evaluated in 10 mice per group. Labeling index 

of bone marrow B lymphocytes was determined by BrdU incorporation. To minimize the 

potential for recovery, mice were sacrificed and sampled within 2 hours of the end of the last 

exposure. The spleen, thymus, and bone marrow data sets were individually evaluated with a 

univariate two-way ANOVA using contrasts to determine treatment effect within each time 

point. Statistical significance was determined at a level of p<0.05. 

Exposure to 100 or 200 ppm (319 or 639 mg/m3

) benzene induced rapid and persistent 

reductions in femoral B, splenic T and B, and thymic T lymphocytes; total nucleated thymus 

72
cells count; and total nucleated spleen cells. There was a significant decrease in thymic cell 

count and B lymphocytes in the spleen at week 2 in mice exposed to 10 ppm (32 mg/m3

) in the 

8-week study and no effects at 1 ppm (3.2 mg/m3). These parameters, however, were 

comparable with the controls by week 4 of the 8-week exposure experiment, and in the separate 

4-week experiment there were no significant decreases in these parameters at any time in the 

mice exposed to 10 ppm (32 mg/m3

) benzene or less. Thus, the 100 ppm (319 mg/m3) level was 

considered a LOAEL and 10 ppm (32 mg/m3) a NOAEL for these effects. The percentage of 

femoral B lymphocytes and thymic T lymphocytes in apoptosis was increased 6- to 15-fold by 

200 ppm (639 mg/m3

). Replication of femoral B lymphocytes was increased during the exposure 

period in the bone marrow, apparently in compensation for the lymphocyte loss induced by 100 

or 200 ppm (319 or 639 mg/m3) benzene exposure. 

Robinson and coworkers (1997) exposed male Sprague-Dawley rats (16/group) to 0, 30, 

200, or 400 ppm (0, 96, 639, or 1278 mg/m3

) benzene 6 hours/day, 5 days/week for either 2 or 4 

weeks. Of the 16 animals in each group, 8 were challenged intravenously with sheep red blood 

cells (SRBCs) at 4 days before the end of the exposure to allow an assessment of humoral 

immunocompetence using the enzyme-linked immunosorbent assay technique. Thymus and 

spleen weights were determined, as were total cell counts for the spleen and femur. Spleen 

CD4+/CD5+, CD8-/CD5+, and Kappa+ lymphocytes were determined by immunostaining. 

Statistical evaluations were performed using Dunnett’s test. 

Total spleen cell counts were significantly reduced (29%) in animals exposed to 400 ppm 

(1278 mg/m3) after 4 weeks, and thymus weight was also significantly reduced (28%). The 

spleen weight and cellularity were comparable with the controls at both 2 and 4 weeks in the 200 

ppm (639 mg/m3) exposure group. After 4 weeks at 400 ppm (1278 mg/m3) there was a 

significant reduction in spleen B, CD4+/CD5-, and CD5+ T lymphocytes. Rats exposed to 30, 

200, or 400 ppm (99, 639, or 1278 mg/m3

) benzene for 2 or 4 weeks and then challenged with 

SRBCs developed a humoral response comparable to the controls’, and only rats exposed to 400 

ppm (1278 mg/m3

) for 2 weeks showed a significant reduction in spleen B lymphocytes 

(Robinson et al., 1997). These results indicate that 400 ppm (1278 mg/m3

) is a LOAEL and 200 

ppm (639 mg/m3) is a NOAEL for immunotoxicity in rats. 

Taken as a group, the results demonstrate that the reduction in bone marrow function 

following benzene exposure is related to decreases in the growth and maturation of pluripotential 

stem cells (CFU-S), lineage-restricted stem cells (e.g., GM-CFU), progenitor cells in various 

stages of maturation, and the stromal cells that provide growth factors necessary for bone 

marrow function. These studies tend to demonstrate the sensitivity of erythropoiesis and 

lymphopoiesis to benzene and the relative resistance of granulopoiesis to benzene. 

Snyder and Kalf (1994) reviewed the effects of benzene on bone marrow cells. Bone 

marrow stromal cells, which include leukocytes, erythrocytes, endothelial cells, reticular cells, 

and fat cells, constitute the hematopoietic connective tissue in which stem cells undergo 

maturation and amplification to form the cells of the circulation. Garnett et al. (1983) reported 

that the marrow adherent layer from benzene-treated animals had an altered morphology and 

failed to support the differentiation of stem cells. In a related study, Kalf et al. (1996) 

demonstrated that p-benzoquinone, an important reactive metabolite of benzene, prevents the 

conversion of pre-interleukins-1 and -1 to the active cytokines by inhibiting calpain and the 

73
interleukin-1 converting enzyme, respectively. Thus, a key benzene metabolite inhibits the 

formation and release of an important cytokine, IL-1, from stromal macrophages. 

Irons et al. (1992) isolated human CD34+ human bone marrow myeloid progenitor cells, 

which proliferate in response to IL-3 and GM-CSF. Addition of hydroquinone, within a limited 

range of concentrations, resulted in synergistic enhancement of colony formation. Hazel et al. 

(1996) made similar observations using mouse bone marrow CD32 cells. The mechanism is 

thought to include the up-regulation of the 5-lipoxygenase pathway in these cells, with increases 

in the formation of LTD4, which in turn reacts with the LTD4 receptor to increase cell 

replication. Hydroquinone independently reacts with the LTD4 receptor to also enhance cell 

replication. Morphological evaluation suggests that maturation of the myeloblast stage to the 

myelocyte stage has been enhanced. These cells remain in their immature state in the presence 

of hydroquinone and do not advance to mature neutrophils, but proliferation continues. 

The implications of these observations are significant in terms of a mechanism by which 

benzene can cause expansion of the pool of myelocytes, which do not then undergo either further 

differentiation to neutrophils or apoptosis, which is the usual method used by the bone marrow to 

control the size of the myelocyte pool. This might provide a pathway for a benzene-induced 

leukemogenic response. It may be hypothesized that if benzene, acting via DNA adduct 

formation or oxidative damage to DNA, causes cell transformation at an earlier stage of 

differentiation, the creation of an expanding pool of myelocytes by benzene and hydroquinone 

provides a mechanism for promotion. Thus, the criteria for a two-stage mechanism of 

carcinogenesis (i.e., initiation followed by promotion) are satisfied. However, the specific issue 

of progression has not been addressed by studies of this type. 

The studies by Tennant et al. (1995) have demonstrated the production of a granulocytic 

leukemia in TgAC mice when treated with benzene dermally but not when given benzene by 

gavage. This is the first animal model that we have had for a benzene-induced granulocytic 

leukemia. The TgAC mouse is derived from the FVB mouse in which the viral Ha-ras gene has 

been inserted into chromosome 8. The granulocytic leukemia develops within 6 months. TgAC 

mice may represent initiated animals in which benzene and hydroquinone act as promoting 

agents leading to proliferation of the leukemia cells, and thus could support the argument for the 

two-stage model for benzene-induced leukemogenesis. It is significant that in the course of 

these studies bone marrow depression was observed, indicating the close relationship between 

noncancer and cancer endpoints in benzene exposure. 

The following discussion offers some thoughts about the attempt to understand the 

mechanism by which DNA damage occurs as a result of benzene exposure. DNA damage is not 

extensively dealt with here because this document concentrates on noncancer endpoints. It is not 

clear, however, that there is no impact on noncancer endpoints by DNA damage. Some 

discussion of DNA damage leading to translocation and gross breaks as well as a discussion of 

DNA repair and the impact of benzene metabolites is warranted. The study by Snyder and Kalf 

(1994) and the more recent study by Smith et al. (1998) relate to benzene-induced chromosome 

damage. On the issue of DNA damage induced by benzene, there is ample evidence to show that 

benzene metabolites covalently bind to protein and DNA. Also, the structures of putative DNA 

adducts are known. An alternative mechanism for DNA damage is via the generation of reactive 

oxygen species. For example, p-benzoquinone and hydroquinone increase superoxide, nitric 

74
oxide, and hydrogen peroxide in HL-60 cells activated with phorbol ester. Boersma et al. (1994) 

suggested that hydroquinone itself is unlikely to generate reactive oxygen species. 

An interesting complimentary alternative mechanism, proposed by Singh et al. (1994), 

suggests that benzene- induced DNA damage is mediated by (1) release of free iron in the bone 

marrow of benzene-treated animals (probably by polyphenolic metabolites of benzene), followed 

by (2) the chelation of iron by hydroquinone or benzenetriol to yield (3) a reactive oxygengenerating species such as superoxide, which in turn (4) causes oxidative damage to DNA. 

Their suggestion that glutathionyl-hydroquinone may be a key intermediate resonates well with 

the suggestion of Brunmark and Cadenas (1988). In a radical-rich system, glutathionylhydroquinone may well be converted to glutathionyl-benzenetriol. Singh et al. (1994) suggest a 

chemical structure for the iron-benzenetriol chelate in which the iron is bound between the 

hydroxyl group at positions 1 and 2 on 1,2,4-benzenetriol with the release of two protons. 

Indeed, it may well be that the auto-oxidation of glutathionyl-benzenetriol is enhanced with an 

iron chelated between positions 1 and 2. These hypotheses can be reconciled by assuming that 

the superoxide produced by the auto-oxidation of gluthionyl-benzenetriol is converted to 

hydroxyl radical and singlet oxygen by the benzenetriol-iron chelate. The resulting reactive 

oxygen species can hydroxylate guanine and other DNA bases, resulting in a mutagenic effect. 

The discussion of the mechanisms of benzene toxicity provides a framework for 

understanding the total disease process. It has become clear that benzene 

toxicity/leukemogenesis is a continuity of effects. Although this is not intended to be a 

discussion of the carcinogenic effects of benzene, all of the effects on the bone marrow caused 

by benzene appear to be time- and dose-related in a sequential manner. The literature suggests 

that chronic low doses lead to cytopenias, which may be reversible if exposure ceases. At 

intermediate chronic doses cytopenias persist and may eventually lead to myelodysplastic 

syndrome (MDS), which is a preleukemic state terminating in AML. If the cytopenias are 

sufficiently severe the patient may succumb to an infection due to failure of the immune system. 

Higher doses may lead directly to bone marrow aplasia and death without the characteristic 

appearance of MDS. Death is usually the result of immune deficiency. Thus, the cytopenias (the 

ultimate form of which is aplastic anemia) leading to decreases in circulating cell numbers and 

immune impairment are not separable from MDS and, eventually, AML. 

The following studies of exposure to benzene and its metabolites are considered briefly 

because of their value in the study of the mechanisms of benzene toxicity. 

Tunek et al. (1982) examined the hematotoxicity of benzene and two of its phenolic 

metabolites, hydroquinone and catechol, in male NMRI mice. Daily doses of 440 mg/kg 

benzene to body weight were injected subcutaneously for 6 days, and CFU-C/tibia, cells/tibia, 

and micronuclei/2000 PCE were assayed on the day after the last injection. Micronuclei 

appeared after one injection, peaked after three injections, then decreased with injections 4, 5, 

and 6. The investigators presumed that the numbers of micronuclei decreased as a result of 

toxicity and cell death before the cells reached the PCE stage. By day 4, the CFU-C/tibia and 

number of cells/tibia were reduced to about 5–10% of control values. 

In another experiment, the animals received doses ranging from 0.7 to 440 mg/kg/day 

benzene on 6 consecutive days and were sacrificed on day 7. Increased levels of micronuclei 

75
were observed at doses above 28 mg/kg/day, and their frequency increased dramatically at doses 

above 55 mg/kg/day. At ~100 mg/kg/day, micronuclei levels were about 50/2000 PCE, whereas 

at the highest dose of 440 mg/kg/day, they were 218/2000 PCE (vs. 3 micronuclei/2000 PCE for 

controls). CFU-C/tibia and bone marrow cellularity were suppressed by about 40% at doses as 

low as 3.5 mg/kg/day. At 100–440 mg/kg/day, both values dropped to 5–10% of control values. 

Six injected doses of hydroquinone ranging from 20 to 100 mg/kg produced micronuclei, 

with a sharp increase occurring at 6 × 80 mg/kg. Bone marrow cellularity was slightly elevated 

at low doses and was suppressed at higher doses, with the number of CFU-C following a similar 

pattern. Catechol, injected at doses ranging from 5 to 42 mg/kg for 6 consecutive days, had no 

effect on any of the three parameters examined. The toxic effects of benzene were alleviated by 

~30–40% when toluene (which competitively inhibits benzene metabolism) was injected 

simultaneously. The toxic effects of hydroquinone were somewhat reduced by simultaneous 

administration of toluene, but not nearly to the extent observed with benzene. The investigators 

calculated that a dose of 440 mg/kg benzene would yield an excretion of ~20 mg/kg 

hydroquinone. However, 20 mg/kg hydroquinone was virtually nontoxic, whereas 440 mg/kg 

benzene produced severe hematologic effects. On the basis of these differences and the 

differences in the sequence of the responses to the two chemicals, Tunek et al. (1981) speculated 

that the hematotoxicity of benzene at low doses is due to agents other than hydroquinone, but 

that effects at and above the threshold dose result from the metabolic formation and 

accumulation of hydroquinone. 

Using an iron-uptake method, Snyder et al. (1989) demonstrated that i.p. administration 

of benzene, hydroquinone, p-benzoquinone, and MUC, singly and in combination with each 

other, inhibited erythropoiesis in female mice. The combination of hydroquinone plus MUC was 

most effective in decreasing iron uptake. Toluene alleviated the effects of benzene but not the 

effects of hydroquinone or p-benzoquinone. 

MacEachern et al. (1992) evaluated the effects of benzene on the morphology and 

function of bone marrow phagocytes. Male Balb/c mice received i.p. injections of either 

benzene (660 mg/kg) or a combination of hydroquinone and phenol (50 mg/kg each) once per 

day for 3 days. Control animals received corn oil or phosphate-buffered saline. The animals 

were sacrificed after the last injection and the following assays were performed: characterization 

and quantitation of subpopulations of bone marrow cells using monoclonal antibody techniques, 

measurement of chemotaxis (the process by which activated phagocytes migrate to an injured 

site in a tissue), and measurement of the oxidative metabolism of phagocytes (to evaluate 

maturation and activation of phagocytes). The bone marrow contained three distinct populations 

of cells: population 1—a larger more dense population (41%); population 2—a population of 

intermediate size and density (23%); and population 3—a smaller, less dense population (33%). 

Population 1 consisted of 85–90% granulocytes, including neutrophils, basophils, and 

eosinophils; population 2 contained a mixture of mononuclear phagocytes (35–40%) and 

immature precursor cells (55–65%); and population 3 contained lymphocytes (86%) and 

immature precursor cells (14%). This pattern of distribution is similar to that observed for 

human peripheral blood leukocytes and bone marrow cells (Landay and Bauer, 1988; LundJohansen et al., 1990, both cited in MacEachern et al., 1992). Differential staining of sorted 

mononuclear phagocytes revealed an increase in the number of mature, morphologically 

activated macrophages in the bone marrow of benzene-treated mice. 

76
Granulocytes and mature macrophages were the only cells to migrate toward 

chemoattractants C5a or TPA, respectively, and cells from benzene-treated mice exhibited 

increased chemotactic activity when compared with controls (p0.02). The migration of bone 

marrow granulocytes from mice treated with hydroquinone and phenol toward C5a was 

depressed (p0.02), whereas the migration of the macrophages from these mice toward TPA was 

enhanced (p<0.02). 

Benzene treatment also induced a significant increase in basal oxidative metabolism in 

bone marrow granulocytes (population 1), but treatment with phenol and hydroquinone did not. 

The investigators concluded that phagocytes and granulocytes from the bone marrow of 

benzene-treated mice are activated and stimulated to differentiate. Further studies are in 

progress to determine whether these phagocytes or their mediators contribute to the development 

of benzene hematotoxicity. 

Gaido and Wierda (1984) demonstrated that in vitro treatment with hydroquinone and pbenzoquinone decreased the ability of stromal cells to support granulocyte/macrophage colony 

formation. Catechol and 1,2,4-benzotriol inhibited colony formation, but only at very high 

concentrations. 

Eastmond et al. (1987) demonstrated that coadministration of 75 mg/kg phenol and 

25–75 mg/kg hydroquinone 2 times/day for 12 days i.p. to mice resulted in a significant, doserelated depression in bone marrow cellularity. The compounds alone produced neither a 

significant nor a dose-related response. Catechol had no effect, either alone or in combination. 

Hydroquinone (10 mg/kg/day), benzoquinone (2 mg/kg/day), or benzenetriol (6.25 

mg/kg/day) administered i.p. to rats produced significant decreases in bone marrow cell counts, 

RBCs, and Hgb (Rao et al., 1988). Trans,trans-muconaldehyde administered to CD-1 mice i.p. 

daily for 10 and 16 days produced hematotoxicity similar to that of benzene (Witz et al., 1985). 

Benzene metabolites added to macrophage cultures significantly and selectively inhibited 

macrophage function (Lewis et al., 1988). 

Irons et al. (1992) observed an enhanced colony-forming response by mouse bone 

marrow cells treated in vitro, first with hydroquinone and then with recombinant 

granulocyte/macrophage colony-stimulating factor. However, treatment with phenol, catechol, 

or MUC plus the colony-stimulating factor did not enhance colony formation. The combination 

of hydroquinone and the stimulating factor also appeared to recruit a segment of the myeloid 

progenitor cell population that was not normally responsive to the stimulating factor. These 

alterations in the differentiation of the myeloid progenitor cell population may be relevant in the 

pathogenesis of chemically induced AML. 

4.2.1.6. Summary of Principal Hematotoxic Effects 

A considerable number of studies in experimental animals have pointed to the changes in 

peripheral blood and bone marrow induced by benzene as being among the most sensitive effects 

of the compound’s toxicity. That these effects were seen in most, if not all, of the species tested, 

as well as in humans occupationally exposed to benzene (see Section 4.1.2.1.), suggests that their 

threshold dosimetry may be important in identifying a point of departure for the derivation of 

77
quantitative risk estimates. Principal among the adverse effects described are leukopenia, 

lymphocytopenia, granulocytosis, anemia, and reticulocytosis. Comparing the incidence of one 

response with another, lymphocyte counts appear to be depressed sooner and more severely than 

other cell types, and granulocytes may be the most resistant (Snyder et al., 1978, 1980). 

Differences in responsiveness to benzene have been observed among species, with mice being 

more sensitive than rats to the hematotoxicity of benzene (Ward et al., 1985; Snyder et al., 1978, 

1980). Intraspecies variability was demonstrated by Snyder et al. (1978, 1980), who reported 

that AKR/J mice were more sensitive than C57BL/6 mice to the compound. Ward et al. (1985) 

showed that in CD-1 mice, males were more sensitive than females. 

A striking feature of the studies of Ward et al. (1985) and Cronkite et al. (1985) was that 

hematologic abnormalities were evident after only 2 weeks of exposure, with comparatively little 

progression during the rest of the exposure period. However, the results of Cronkite et al. (1985, 

1989) point to the possibility that peripheral blood elements may have the capacity to recover 

from benzene-induced toxicity, depending on the duration of exposure and concentration. For 

example, lymphocyte counts, which were depressed in C57BL/6 mice exposed to benzene at 300 

ppm (32–1279 mg/m3

) for 2–16 weeks, tended toward normal values during a 4- to 8-week 

recovery period. 

Dose-response studies such as that of Ward et al. (1985) have served as a basis for the 

identification of NOAELs and LOAELs. For example, a NOAEL of 30 ppm (96 mg/m3

) and a 

LOAEL of 300 ppm (958 mg/m3

) have emerged for CD-1 mice from the Ward et al. study. The 

data from this study have also served as a basis for BMD modeling. Using the BMDL derived 

from this modeling as a point of departure for RfC determination yields a value that is in good 

agreement with the RfC obtained from the hematotoxicity data from the occupational exposure 

study of Rothman et al. (1996a), as described in Section 5.1. 

Benzene-induced peripheral blood abnormalities reflect a disruption of all levels of 

hematopoiesis in the bone marrow. For example, several studies have demonstrated 

concentration-dependent depression of bone marrow cellularity in mice treated with benzene, the 

lowest concentration producing this effect being 21 ppm administered for 2 weeks (Toft et al., 

1982). Similarly, the stem cell compartment (CFU-S) appears to be sensitive to the adverse 

effects of benzene, particularly when drawn into the cell cycle in response to benzene-induced 

damage to other cells (Snyder, 1987). Thus, concentrations of 100 ppm (320 mg/m3

) benzene 

have produced significant depressions in the CFU-S population in the bone marrow of mice 

exposed for 2–4 weeks (Cronkite et al., 1985; Seidel et al., 1989). 

The granulocytic and erythropoiesis progenitor cells are also sensitive to the effects of 

the compound, exhibiting significant depression at 21 ppm (67 mg/m3

) benzene for 2 weeks and 

10 ppm (32 mg/m3) for 4 weeks, respectively (Toft et al., 1982; Seidel et al., 1989). In addition, 

both the stem cells and the progenitor cells can recover from benzene-induced depression, at 

least partially, depending on the concentration and duration of exposure. Cronkite et al. (1989) 

observed that depression of CFU-E and BFU-E were paralleled by depressions in peripheral 

RBC counts. 

Bone marrow macrophages also appear to be a target of benzene toxicity (MacEachern et 

al., 1992). Thus, stromal macrophages from benzene- or hydroquinone-treated mice failed to 

78
convert IL-1 to the mature cytokine 17-Kd, a component essential for hematopoiesis (Renz and 

Kalf, 1991). This allows the suggestion that the inhibitory effect of benzene on lymphocyte 

proliferation can be mediated through the suppression of cytokine production. 

Another important consideration is whether the benzene-induced perturbation of the 

hematopoietic system is mediated through the effects of the parent compound or via one or more 

metabolites. The fact that both in vivo and in vitro studies have demonstrated the ability of some 

metabolites to induce qualitatively similar toxicological effects to those of benzene provides 

evidence consistent with the concept that the hematotoxicity of the compound may be mediated 

through its metabolites. 

4.2.2. Reproductive/Developmental Effects 

4.2.2.1. Reproductive Toxicity 

Multiple-generation reproductive toxicity studies of benzene were not found in the 

literature. Relevant information pertinent to the reproductive toxicity of benzene in animals is 

described below. 

4.2.2.1.1. Oral exposure. NTP conducted the 90-day subchronic toxicity and 2-year bioassay 

studies of benzene in mice and rats (NTP, 1986). Other than ovarian effects, no signs of 

reproductive toxicity were reported. Benzene was administered by gavage to B6C3F1 mice and 

F344/N rats. In the 90-day study, the doses of benzene, administered in corn oil 5 days/week, 

ranged from 25 to 600 mg/kg; in the 2-year study, doses were 50, 100, or 200 mg/kg for male 

rats and 25, 50, or 100 mg/kg for female rats and all mice. There were no ovarian effects in 

either the rats or the mice treated for 90 days or in the rats treated for 2 years. Mice treated for 2 

years exhibited ovarian lesions ranging from atrophy to neoplasia. However, the incidence of 

nonneoplastic lesions was not dose related. The nonneoplastic lesions and their incidence, based 

on the total number of animals examined for each group, for mice administered 0, 50, 100, or 

200 mg/kg benzene were ovarian atrophy (32, 79, 65, and 46%, respectively) and epithelial 

hyperplasia (26, 89, 63, and 60%, respectively). 

Spano et al. (1989) examined the cytotoxic action of benzene on mouse germ cells using 

flow cytometric DNA content measurements. Testicular monocellular suspensions were 

obtained from (C57BL/Cne × C3H/Cne) F1 male mice receiving single doses by gavage of 0, 1, 

2, 4, 6, or 7 mL/kg benzene to body weight. The effects were measured in three animals per 

group 7, 14, 21, 28, and 70 days after treatment. Testicular cells were classified as mature 

haploid, immature haploid (haploid is split into two parts because of different staining on the 

elongated and round spermatids), diploid, or tetraploid, depending on DNA content. 

Benzene treatment did not affect body weight or testes weight, but it did alter the ratio of 

testicular cell types. DNA histograms of mouse testis cells obtained at different times after 

benzene exposures showed a dose-related decrease in the tetraploid cell fraction (mainly primary 

spermatocytes); 7 days after treatment the tetraploid cell number in animals exposed to 6 and 7 

mL/kg benzene was depressed to ~80% of control values. The percentage of round spermatids 

in animals exposed to 7 mL/kg benzene was also decreased to ~80% of control values; however, 

in this case the dose relationship was not distinct. These effects indicate cytotoxicity of 

79
differentiating spermatogonia. Dose-dependent recovery processes for the tetraploid cells had 

begun 21 days after treatment, simultaneously with reduction of other cell subpopulations, but 

they were still incomplete by 70 days. Recovery of the round spermatids began by day 28 

posttreatment and was complete by day 70. There was also a time- and exposure-dependent 

reduction (to 60% of control values for the 4, 6, and 7 mL/kg doses) in the percentage of 

elongated spermatids at 28 days after treatment; recovery was complete by 70 days. The 

investigators concluded from this study that benzene can induce acute cytotoxicity in mouse 

germ cells. Both the NTP (1986) and the Spano et al. (1989) studies provide valuable 

descriptive information regarding the effects of benzene on the reproductive system. It should 

be noted that longer term and more relevant studies are needed to confirm the conclusion of 

Spano et al. (1989). 

4.2.2.1.2. Inhalation exposure. The available studies on the reproductive toxicity of inhaled 

benzene to test animal species is summarized in Table 7. 

Subchronic toxicity studies have provided information relevant to the reproductive 

toxicity of benzene. In an early inhalation study, Wolf et al. (1956) demonstrated moderately 

increased testicular weights in groups of 10–25 rats exposed to 6600 ppm (21,084 mg/m3

) 

benzene for 13 weeks but not in rats exposed to 88 ppm (281 mg/m3

) for 30 weeks, to 2200 ppm 

(7028 mg/m3

) for 30 weeks, to 4400 ppm (14,056 mg/m3) for 5 weeks, or to 9400 ppm (30,030 

mg/m3

) for more than 1–19 days. Groups of 5–10 male guinea pigs exposed to 88 ppm (281 

mg/m3) benzene for ~9.6 months had a slight increase in testicular weights, but guinea pigs 

Table 7. Reproductive toxicity of inhaled benzene in test animals 

Species 

No./ 

group Treatment Reproductive effects 

Male rat 10–25 21,084 mg/m3

, 7 

hrs/day, 5 days/wk 

for 13 weeks 

Moderate changes in testes 

weight (no quantitative 

data) 

Rabbit 

Male 

guinea pig 

Male and 

female 

SpragueDawley 

rats 

LOAEL/ 

NOAEL 

(mg/m3

) Reference 

LOAEL: NA 

NOAEL: NA 

Wolf et al., 

1956 

LOAEL: NA 

NOAEL: NA 

Wolf et al., 

1956 

LOAEL: NA 

NOAEL: NA 

Wolf et al., 

1956 

256 mg/m3

, 7 

hrs/day, 5 day/wk 

for ~8.5 months 

5–10 281 mg/m3

, 7 

hrs/day, 5 days/wk 

for ~9.6 months 

50/sex 0, 3.2, 32, 96, 958 

mg/m3

 6 hrs/day, 5 

days/wk for 13 

weeks; 

10/sex/group 

sacrificed after 7, 

14, 28, 56, and 91 

days of treatment 

1–2 Slight change in 

histopathology of the testes 

(no details) 

Slight change in weight of 

testes (no quantitative data) 

No microscopic lesions of 

ovaries and testes 

LOAEL: 958a 

NOAEL: 96 

Ward et al., 

1985 

80
Species 

No./ 

group Treatment Reproductive effects 

LOAEL/ 

NOAEL 

(mg/m3) Reference 

CD-1 

mouse 

150/sex 0, 3.2, 32, 96, 958 

mg/m3

 6 hrs/day, 5 

days/wk for 13 

weeks; 

30/sex/group 

sacrificed after 7, 

14, 28, 56, and 91 

days of treatment 

Cystic ovaries (4/10), 

testicular atrophy (7/10), 

decreased sperm count 

(6/10), and increased 

abnormal sperm forms 

(9/10) only at 958 mg/m3 

for 91 days 

LOAEL: 958 

NOAEL: 96 

Ward et al., 

1985 

Female 

SpragueDawley rat 

26 0, 3.2, 32, 96, 958 

mg/m3

 6 hrs/day, 5 

days/wk for a 10-

week period before 

and during mating, 

during gestation to 

GD 20, and from 

day 5 to day 21 of 

lactation 

Reduction in body and liver 

weight in the female pups at 

958 mg/m3

 (p<0.05); no 

effect on maternal 

mortality, body weight, 

physical parameters, 

pregnancy rate, length of 

gestation, number of live 

and dead pups at birth, and 

sex distribution data; no 

effect on pup survival and 

growth, and gross 

postmortem manifestations 

LOAEL: 958 

NOAEL: 96 

Kuna et al., 

1992 

a

Effects on hematologic parameters place the LOAEL for the overall study in rats at 958 mg/m3 and the NOAEL at

96 mg/m3

.

GD = gestation day.

NA = not applicable (only one concentration tested).

81
exposed to 88 ppm (281 mg/m3

) for 4 weeks did not. Groups of 1–2 rabbits exposed to benzene 

concentrations of 88 ppm (281 mg/m3) for ~8.5 months had slight histopathologic changes in the 

testes, described by the authors as degeneration of the seminiferous tubules (no other 

concentrations were tested in rabbits). Exposures were routinely 7 to 8 hours/day, 5 days/week. 

Air-exposed or unexposed animals served as controls. The vapor concentrations were 

maintained within 10% of the desired concentration. No further details were given for the results 

of this study, and quantitation was not possible. Although the study was not specifically 

designed to detect reproductive effects, the gonadal effects noted stimulated further interest in 

the potential reproductive effects of benzene. 

Ward et al. (1985) exposed male and female CD-1 mice and Sprague-Dawley rats to 

benzene concentrations of 1, 10, 30, or 300 ppm (3.2, 32, 96, or 958 mg/m3

) 6 hours/day, 5 

days/week for 13 weeks. Animals exposed to filtered air served as controls. Groups of 10 males 

and 10 females per species were sacrificed after 7, 14, 28, 56, and 91 days of treatment. 

Endpoints of the investigation that were relevant to reproductive toxicity included body and 

organ weights and gross and microscopic pathology. The mice exposed to 300 ppm (958 mg/m3

) 

benzene exhibited statistically significant decreases in testes weights at days 28, 56, and 91 and 

decreased testes/body weight ratios on days 56 and 91 (data not given). Testicular weight for the 

rats was comparable with control values. The mice exposed to 300 ppm (958 mg/m3

) for 91 days 

had testicular lesions that included minimal to moderately severe bilateral atrophy and 

degeneration (7/10 mice), moderate to moderately severe decreases in spermatozoa in the 

epididymal ducts (6/10 mice), and minimal to moderate increases in abnormal sperm forms 

(9/10). Four of 10 female mice had bilateral ovarian cysts after 91 days of exposure to 300 ppm 

(958 mg/m3) benzene. These testicular and ovarian lesions were not observed at earlier 

sacrifices, but similar lesions did appear at lower concentrations; however, biological 

significance is questionable. Microscopic findings in the control animals were not mentioned. 

The rats had no signs of reproductive toxicity; however, hematotoxicity occurred at 300 

ppm (958 mg/m3), giving the overall study a LOAEL of 300 ppm (958 mg/m3) and a NOAEL of 

30 ppm (96 mg/m3

). The experiment appeared to be a carefully performed subchronic toxicity 

study, and the appearance of microscopic changes in the testes and ovaries of the 300 ppm (958 

mg/m3) group and similar—but only occasional—findings at 10 and 30 ppm groups suggest a 

concentration response. 

Kuna et al. (1992) assessed the effects of benzene on female fertility in Sprague-Dawley 

rats. The study tested occupational exposure levels as well as higher exposures that previously 

demonstrated developmental toxicity. Groups of 26 Sprague-Dawley female rats inhaled 

benzene vapor concentrations of 0, 1, 30, and 300 ppm (0, 3.2, 96, and 958 mg/m3) benzene 

(purity, 99.96%) 6 hours/day, 5 days/week for a 10-week premating and mating period, and daily 

from gestation days (GDs) 0 to 20 and lactation days 5 to 20. Daily vaginal smears were 

examined to determine whether estrus was affected by treatment. 

Neither strain had any treatment-related effects, as evidenced by data on maternal 

mortality, body weight, physical parameters, pregnancy rate, length of gestation, number of live 

and dead pups at birth, and sex distribution. No effect was observed on pup survival and growth 

or on gross postmortem findings. In the Sprague-Dawley rats there was a trend toward reduced 

body and organ weights in the 21-day-old pups exposed to 30 and 300 ppm (96 and 958 mg/m3

) 

82
levels; however, the differences were statistically significant (p<0.05) only for reduced body 

weight (89% of control value) and liver weight (80% of control value) in the female pups at 300 

ppm (958 mg/m3

). Increases in relative kidney weight for the female Sprague-Dawley pups were 

attributed to differences in body weight. The CD rats exhibited no such changes. The 

investigators concluded that benzene at concentrations as high as 300 ppm (958 mg/m3) did not 

induce reproductive toxicity in CD rats. The study in CD rats does not identify a LOAEL or a 

NOAEL. The study in Sprague-Dawley rats identifies a LOAEL of 300 ppm (958 mg/m3

) and a 

NOAEL of 30 ppm (96 mg/m3

), based on the statistically significant reduction in body and liver 

weights, and based on past studies showing similar effects at similar concentrations of benzene. 

4.2.2.2. Developmental Toxicity 

4.2.2.2.1. Oral exposure. Seidenberg et al. (1986) evaluated benzene, along with 54 other 

chemicals of known and unknown developmental toxicity potential, to validate the 

Chernoff/Kavlock assay as a screen for developmental effects. Benzene doses of 1300 

mg/kg/day were administered in cottonseed oil by gavage to pregnant ICR/SIM mice on GDs 8 

through 12, and the dams were allowed to deliver. Maternal toxicity was assessed using body 

weights, mortality, or other clinical signs of overt toxicity as endpoints. Benzene had no effect 

on maternal body weight, but it did produce significantly lower neonatal body weight (to ~95% 

of control value, p<0.05), as measured on days 1 and 2 after birth. This was a screening study 

that used only one dose, and therefore it does not provide dose-response information. However, 

the study does demonstrate evidence for fetal toxicity in the absence of observable maternal 

toxicity. 

In the Exxon Chemical Company (1986) study, pregnant female Sprague-Dawley rats 

(20–22/group) were treated by gavage with 0, 50, 250, 500, or 1000 mg/kg/day benzene on GDs 

6–15. No dose-related mortality was observed. Significant findings in the treated dams as 

compared with controls were decreased food consumption at 250 mg/kg, decreased body 

weight, body weight gains at 500 mg/kg, and increased incidence of alopecia at 1000 mg/kg 

(p<0.05). Developmental toxicity was limited to decreased fetal body weights at 500 mg/kg 

(p<0.05). Fetuses were examined only for external malformations, not for skeletal and visceral 

malformations. This study identified a NOAEL of 50 mg/kg/day and a LOAEL of 250 

mg/kg/day for maternal toxicity and a NOAEL of 250 mg/kg/day and a LOAEL of 500 

mg/kg/day for developmental toxicity in Sprague-Dawley rats. 

4.2.2.2.2. Inhalation exposure. This section summarizes the data for the developmental toxicity 

of inhaled benzene in animals. The studies included experiments with three species (rats, mice, 

and rabbits) and intermittent (6 or 7 hours/day) and continuous (24 hours/day) exposures. Some 

studies tested more than one concentration of benzene and demonstrated a concentration 

response and a LOAEL and/or a NOAEL. Some studies used only one concentration and were 

not useful for determining dose-response relationships, but they did provide supporting evidence 

for fetal toxicity. Three studies evaluated fetal hematologic parameters, and one study examined 

the possible synergistic effects of benzene with other chemicals. The available studies on the 

developmental toxicity of inhaled benzene are summarized in Table 8. 

83
Table 8. Developmental toxicity of inhaled benzene in test animals 

Strain/specie 

s 

No. of 

dams/group Exposurea 

Effects Effect levels 

maternal/developmental 

(mg/m3) Reference Maternal Developmental 

Sprague- 14–15 0, 32, 160, or 1597 Decreased body Decreased mean live body LOAEL: 160/160 Kuna and 

Dawley rat exposed, mg/m3, 7 hrs/day on weight and body weight at 160 and 1597 NOAEL: 32/32 Kapp, 1981 

11 controls GDs 6–15; sacrificed weight gain at 160 mg/m3, day 20c; decreased 

on day 20 and 1597 mg/m3, crown-rump distance, 1597 

days 5–15; dose- mg/m3c; skeletal and 

relatedb visceral (brain) 

abnormalities at 160 and 1597 mg/m3; increased incidence of malformations at 1597 mg/m3 

SpragueDawley rat 35–37 exposed, 

32–34 controls 

0, 3.2, 32, 128, or 319 mg/m3, 6 hrs/day 

on GDs 6–15; sacrificed on day 20 

None in any group Decreased body weight at 

319 mg/m3 (p0.05); 

variants in all but one group (including controls), not dose related; no increase in incidence of malformations 

LOAEL: NA/319 NOAEL: NA/128 

Coate et al., 1984 

SpragueDawley rat 14–18 0, 319, 958, or 7028 mg/m3, 6 hrs/day on 

GDs 6–15; sacrificed on day 21 

Decreased body weight gain at 7028 mg/m3 (p< 0.01) 

Decreased body weight and length at 7028 mg/m3 (p< 

0.05); increased skeletal variants all exposure groups (p<0.05 at 319 and 7028 mg/m3; females more 

sensitive than males); no increase in incidence of malformations 

LOAEL: 7028/319 NOAEL: 958/NA 

Green et al., 1978 

Rat 5–12 1.0–670 mg/m3 24 

hrs/day, 10–15 days before mating and 

throughout pregnancy 

No data Tendency toward decreased 

litter sizes at 64 mg/m3; 

complete absence of litters at 670 mg/m3 

LOAEL: ND NOAEL: ND 

Gofmekler, 

1968 

84
Table 8. Developmental toxicity of inhaled benzene in test animals (continued) 

Strain/specie 

s 

No. of 

dams/group Exposurea 

Effects Effect levels 

maternal/developmental 

(mg/m3) Reference Maternal Developmental 

CFY rat 19 exposed, 

28 controls 

0 or 1000 mg/m3, 24 

hrs/day on days 9–14 of pregnancy 

Decreased body weight gain (p<0.01) 

Decreased body weight (p<0.01) and growth retardation (p<0.05); retarded skeletal development and increased incidence of extra ribs and fused sternebrae (p<0.05 for both); no increase in incidence of malformations 

LOAEL: NA NOAEL: NA 

Hudak and Ungvary, 1978 

CFY rat 20–22 

exposed, 48 controls 

0, 150, 450, 1500, or 3000 mg/m3, 24 

hrs/day on GDs 7–14; sacrificed on day 21 

Decreased body weight gain at 150 

mg/m3 (p<0.001), 

somewhat dose related; liver/body weight increased (p<0.05 or 0.01) 

Decreased body weight at 

150 mg/m3 (p<0.001), 

increased resorptions and skeletal and weight retardation (p<0.01–0.05), effects not dose related; no increase in incidence of malformations 

LOAEL: 150/150 NOAEL: NA 

Tatrai et al., 1980 

CFY rat 17 400 mg/m3, 24 

hrs/day on GDs 7–15 

Decreased body weight gain (p<0.001); increased relative liver weight (p<0.05) 

Retarded weight gain (p<0.01); skeletal growth retardation (p<0.001) 

LOAEL: NA NOAEL: NA 

Ungvary, 1985 

85
Table 8. Developmental toxicity of inhaled benzene in test animals (continued) 

Reference 

Keller and Snyder, 1988 

Corti and Snyder, 1996 

Effect levels 

maternal/developmental 

(mg/m3) 

LOAEL: NA/64 

NOAEL: 64/32 

LOAEL: NA NOAEL: NA 

Effects 

Developmental 

16-day-old fetus: no effect 

on hematologic parameters 

2-day-old neonates: reduced circulating erythroid precursor cells (all concentrations) (p<0.05 at 64 mg/m3); increased 

hepatic hemato-poietic blast cells, lympho-cytes, and granulopoietic precursor cells and decreased hepatic erythropoiesis precursor cells (all p<0.05 at 64 mg/m3) 

6-week-old adult: similar 

pattern of enhanced granulopoiesis (64 mg/m3) 

Bone marrow samples from 

6-week-old offspring: protocols (a) and (b) caused changes in CFU-E counts, males only; protocol (c) caused changes in CFU-E counts, females onlyd 

Maternal 

None observed 

No data 

Exposurea 

16, 32, or 64 mg/m3 , 

6 hrs/day on GDs 6–15 

(a) 32 mg/m3 

benzene GDs 6–15 (b) 5% ethanol in drinking water ad lib (c) 32 mg/m3 benzene + 5% ethanol (d) air + distilled water 

No. of 

dams/group 

5 

8 

Strain/species 

Swiss-Webster mouse 

Swiss-Webster mouse 

86
Table 8. Developmental toxicity of inhaled benzene in test animals (continued) 

Strain/species 

No. of 

dams/group Exposurea 

Effects Effect levels 

maternal/developmental 

(mg/m3) Reference Maternal Developmental 

CF-1 mouse 35–37 0 or 1597 mg/m3 

ppm, 7 hr/day on GDs 6–15; sacrificed on day 18 

None Decreased body weight 

(p<0.05), “significantly” increased skeletal variants of fetuses; no increase in incidence of malformations 

but was toxic to fetuses 

LOAEL: NA NOAEL: NA 

Murray et al., 1979 

CFLP mouse 15 exposed, 

115 controls 

0, 500 or 1000 mg/m3, 24 hrs/day on 

GDs 6–15 

Not mentioned Weight and skeletal 

retardation, both concentrations (p< 0.05), somewhat dose-related 

LOAEL: ND/500 NOAEL: NA/NA 

Ungvary and Tatrai, 1985 

New Zealand rabbit 

20 0 or 1597 mg/m3, 7 

hr/day on GDs 6–18; sacrificed on day 29 

None Statistically significant 

decrease in minor skeletal variants, lumbar spurs and proportion with 13 ribs, in exposed fetuses 

LOAEL: NA NOAEL: NA 

Murray et al., 1979 

New Zealand rabbit 

11 or 15 exposed, 60 controls 

0, 500 or 1000 mg/m3, 24 hrs/day on 

GDs 7–20 

Decreased weight gain and increased relative liver weight at 1000 mg/m3 (p< 

0.05) 

Decreased body weight and increased abortions and skeletal variants at 313 ppm (p<0.05 for all effects) 

LOAEL: 1,000/1,000 

NOAEL: 500/500 

Ungvary and Tatrai, 1985 

aConversion factors, 1 ppm = 3.26 mg/m3; 1 mg/m3 = 0.31 ppm bStatistically different from control as determined by pairwise multiple comparison procedures cStatistically different from control as determined by Cochran’s approximation to t (t’) dData from abstract; no other details available GD = gestation day NA = not applicable ND = not determined 

87
In a review of some of the earlier, mostly unpublished developmental toxicity studies on 

benzene, Brief et al. (1980) observed that the developmental effects in animals were 

characterized mainly by fetal toxicity in rats exposed to 40, 50, 500, and 2200 ppm (128, 159, 

1597, and 7028 mg/m3) (Dow Chemical Co., 1992b; Green et al., 1978). Maternal toxicity and 

some fetal malformations occurred at 500 ppm (1597 mg/m3

); concentrations of 10 ppm (32 

mg/m3

) produced conflicting results (Brief et al., 1980). 

Kuna and Kapp (1981) exposed pregnant Sprague-Dawley rats (14–15/group) to benzene 

concentrations of 0, 10, 50, or 500 ppm (0, 32, 160, or 1597 mg/m3

) 7 hours/day on GDs 6–15 

and sacrificed the dams on day 20 to evaluate maternal and developmental effects. Hematologic 

evaluation of the dams, performed on GDs 5 and 20, included RBC, WBC, and differential 

counts. 

There were no deaths, observable illness, or hematologic changes for dams in any 

exposure group. During GDs 5 through 15, maternal body weight gains were significantly 

decreased to 66% of control values in the 50 ppm (160 mg/m3

) group and to 63% of control 

values in the 500 ppm (1597 mg/m3

) group during GDs 5–15 (p<0.05 for both groups), whereas 

during GDs 15–20, body weights in the dams exposed to 10 ppm (32 mg/m3) and weight gain in 

the dams exposed to 10 and 500 ppm (32 and 1597 mg/m3

) were increased. Body weight 

corrected for gravid uterine weight was not determined. No differences were observed between 

exposed and control groups in the number of implantation sites per number of ovarian corpora 

lutea (implantation efficiency); the incidences of resorbed, dead, or live fetuses; and the sex 

distribution. 

Fetal-crown rump lengths were decreased in the 500 ppm (1597 mg/m3

) group, and mean 

body weight of live fetuses were decreased in both the 50 and 500 ppm (160 and 1597 mg/m3) 

groups (control, 4.4 ± 0.6 g; 50 ppm [160 mg/m3

], 3.8 ± 0.7 g; 500 ppm [1597 mg/m3], 3.6 ± 0.8 

g; all statistically significant, p<0.05). One fetus from each of four litters from dams treated with 

500 ppm (1597 mg/m3

) benzene displayed exencephaly (one fetus), angulated ribs (one fetus), or 

ossification of the forefeet out of sequence (two fetuses); these abnormalities were not observed 

in control fetuses. This group also exhibited delayed ossification in the skull, vertebral column, 

rib cage, pelvic girdle, and extremities and significantly fewer tail bones than the controls 

(p<0.05). There was evidence of a dose-related (but not statistically significant) decrease in the 

mean number of phalanges and metacarpals. 

In summary, 13 litters and 142 fetuses were examined from the group exposed to 500 

ppm (1597 mg/m3

); 30 fetuses from six litters had delayed ossification (p<0.05), and four fetuses 

from four litters had skeletal variants and abnormalities. In the group exposed to 50 ppm (160 

mg/m3

), 125 fetuses from 15 litters were examined; 23 fetuses from six litters had variants 

(p<0.05). The investigators concluded that the effects observed in this study were benzene 

induced at concentrations of 50 and 500 ppm (160 and 1597 mg/m3

). The study appears to have 

been conducted according to standard protocols and recommendations (U.S. EPA, 1991) for 

developmental toxicity. The study establishes a LOAEL for the developmental toxicity of 

inhaled benzene of 50 ppm (160 mg/m3) and a NOAEL of 10 ppm (32 mg/m3). 

Coate et al. (1984) exposed pregnant Sprague-Dawley rats by inhalation to 0, 1, 10, 40, 

or 100 ppm (0, 3.2, 32, 128, or 319 mg/m3) benzene 7 hours/day on GDs 6–15. Benzene levels 

88
in the chambers, monitored at least once per day using both infrared and gas chromatography 

methods, were within 10% of target concentrations by both analyses, except for the gas 

chromatography analysis of the 10 ppm (32 mg/m3

) concentration, which was within 17% of 

target. 

Maternal body weight and mortality rate did not differ between treated and control 

animals (other than a slight but insignificant decrease in body weight of the 100 ppm [319 

mg/m3

] group), and there were no clinical signs or gross pathology suggestive of maternal 

toxicity during gestation. Body weight corrected for gravid uterine weight was not determined, 

and hematologic assays were not performed. The average number of implantations, number of 

resorptions, resorption incidence, number of live fetuses, and the incidence of dams with one or 

more resorbed implantations were comparable in all control and exposed groups. The fetuses in 

the 100 ppm (319 mg/m3

) group exhibited a significant decrease (p0.05) in average male and 

female fetal body weights and a statistically insignificant decrease in fetal length. These values 

were within 10% of control values. There were no significant differences in the percentage of 

fetuses per litter with one or more variants, although slight increases in variants (dilation of the 

renal pelvis and ureters) occurred in the 1 and 100 ppm (3.2 and 319 mg/m3

) groups. Soft tissue 

anomalies occurred in all exposed groups, but they were not dose related. Delays in ossification 

of the skull, vertebral centra, and extremities occurred at 100 ppm (319 mg/m3

). The 

investigators concluded that benzene is weakly toxic to the fetus at 100 ppm (319 mg/m3

) (a 

concentration that was not toxic to the dams) and is not toxic to the fetus at 40 ppm (128 mg/m3

). 

This study establishes a LOAEL of 100 ppm (319 mg/m3

) and a NOAEL of 40 ppm (128 mg/m3) 

for the fetal toxicity of benzene. 

Green et al. (1978) exposed groups of pregnant Sprague-Dawley rats to benzene 

concentrations of 100, 300, and 2200 ppm (319, 958, and 7028 mg/m3) 6 hours/day on GDs 

6–15. Daily body weights were measured as an indicator of maternal toxicity; peripheral blood 

cell counts were not performed. The dams were sacrificed on day 21, and developmental 

toxicity was evaluated. Maternal weight gain was similar to control values for the rats exposed 

to 100 and 300 ppm (319 and 958 mg/m3

); however, body weight for the females exposed to 

2200 ppm (7028 mg/m3) was significantly reduced from GDs 8 to 20 to ~44%–83% (p<0.01 for 

all days) of control values. There were no differences between all exposure groups and controls 

with regard to implantation sites/litter, live fetuses/litter, percentage resorption/implantation site, 

percentage litters with resorption, litters totally resorbed, and resorption/litter with resorption. 

At concentrations of 100 and 300 ppm (319 and 958 mg/m3

), fetal sex ratio, mean fetal body 

weight, and mean fetal crown-rump length were comparable with control values; at 2200 ppm 

(7028 mg/m3), mean fetal weight and mean crown-rump length were significantly lower than 

controls (p<0.05). 

The incidence of delayed ossification was similar for the exposed groups and controls; 

however, among the fetuses exposed to 300 and 2200 ppm (958 and 7028 mg/m3

) and exhibiting 

delayed ossification, the incidence was increased significantly for females over males (p<0.05). 

The number of litters displaying an increased incidence of unossified sternebrae was increased at 

100 ppm (319 mg/m3) (9/18 litters) and 2200 ppm (7028 mg/m3) (11/15 litters), and at 2200 ppm 

(7028 mg/m3

) the number of females with the abnormality was significantly increased over 

males (p<0.05). The concentration of 2200 ppm (7028 mg/m3

) benzene was maternally toxic; 

however, fetal toxicity manifested as skeletal abnormalities was observed at concentrations that 

89
were not maternally toxic (and were statistically significant). The apparently increased 

sensitivity of the female fetus to the effects of benzene is supported by other observations that 

female rabbits, mice, and rats are more sensitive to these effects of benzene (Desoille et al., 

1961; Ito, 1962a-d; Sato et al., 1975). The investigators tentatively suggested that the 

differences in the responses of the males and females may be related to hormonal differences. 

In an early study, Gofmekler (1968) exposed female rats to benzene concentrations 

ranging from 0 to 210 ppm (670 mg/m3

) 24 hours/day for 10–15 days prior to mating and all 

during pregnancy. The rats exhibited a dose-related decrease in litter size and produced no 

litters at the highest concentration of 210 ppm (670 mg/m3

). This effect was not observed in the 

experiments of Green et al. (1978) at higher benzene concentrations administered over a 

different exposure period (6 hours/day on GDs 6–15). This suggests that the rat embryo is more 

sensitive to continuous exposure than to intermittent exposure only during organogenesis. 

However, females in that study were exposed before mating, and ovulation rate and fertility may 

have been affected independently of developmental effects. The Gofmekler (1968) study 

demonstrated significant differences in the weight of liver, lung, spleen, kidney, and thymus 

from fetuses of benzene-treated dams. The weights were mostly decreased, but the effect was 

not dose related. Gofmekler (1968) concluded that exposure of pregnant rats to relatively low 

concentrations of benzene can disturb normal fetal development. However, the lack of a 

concentration response in the organ weight data for fetal rats precludes the determination of a 

LOAEL or a NOAEL for the study. 

Hudak and Ungvary (1978) conducted an experiment with CFY rats exposed to 313 ppm 

(1000 mg/m3

) benzene 24 hours/day on GDs 9 through 14. Exposure to benzene resulted in 

statistically significant decreases in maternal body weight gain (p<0.01), fetal weights (p<0.01) 

and percentage of weight-retarded fetuses (p<0.05) and increased incidences of extra ribs and 

fused sternebrae (p<0.05 for both). The use of only one dose precluded the derivation of a 

LOAEL and/or a NOAEL for the study. 

A later study (Tatrai et al., 1980) examined the developmental toxicity of benzene using 

several doses. Pregnant CFY rats inhaled 0, 47, 140, 465, or 930 ppm (0, 150, 450, 1500, or 

3000 mg/m3

) 24 hours/day on GDs 7–14. Animals were sacrificed on day 21 to evaluate 

developmental effects. The dams exhibited significantly decreased body weight gain at all 

concentrations (p<0.001, ANOVA). The effect was dose related at 47, 140, and 465 ppm 

(150,450, and 1500 mg/m3), but not at 930 ppm (3000 mg/m3). For example, the body weight 

gain for dams exposed to 465 ppm (1500 mg/m3) benzene was 28.5% of the starting body weight 

compared with the control value of 62.8%, whereas the body weight gain for dams exposed to 

930 ppm (3000 mg/m3

) benzene was 37.0%. Liver/body weight ratios were significantly 

increased in the dams exposed to 140, 465, or 930 ppm (450, 1500, or 3000 mg/m3

), with a 

response pattern similar to the body weights (p<0.05 or p<0.01, ANOVA). 

Fetal body weights were decreased at 47 ppm (150 mg/m3

) (p<0.001, all 

concentrations); fetal loss in the percent of total implantation sites was increased 4–7 times 

control values at 140 ppm (450 mg/m3

) (p<0.05 or p<0.01, Mann Whitney U test); and skeletal 

and weight retardation were observed in the groups exposed to the three highest concentrations 

(p<0.01 or p<0.05). For these effects, there was no distinct relationship between concentration 

and response, and there was no increase in the incidence of malformations, even at 

90
concentrations 47 ppm (150 mg/m3

) that were maternally toxic. The main difference observed 

for the days 7–14 exposure regimen (Tatrai et al., 1980) study versus the days 9–14 regimen 

(Hudak and Ungvary, 1978) was the increase in fetal mortality (30–40% of the implants) that 

occurred when exposure was started on GD 7. 

To test the hypothesis that the congenital effects of industrial solvents may be attributable 

to simultaneous exposure to a combination of two or more chemicals, Ungvary and Tatrai (1985) 

conducted a study using combinations of benzene and its methyl derivatives, toluene and xylene, 

with each other or with acetylsalicylic acid (ASA), a known developmental toxicant. Groups of 

pregnant CFY rats inhaled combinations of 123 ppm (400 mg/m3) benzene and 307 ppm (1000 

mg/m3

) toluene or 189 ppm (600 mg/m3) xylene 24 hours/day on GDs 7–15. Additional groups 

of 7–14 animals were exposed to 798 ppm (2600 mg/m3

) benzene on GDs 10–12 and given oral 

doses of either 250 or 500 mg/kg of ASA on GD 12. 

At 400 mg/m3

, benzene produced maternal toxicity, as evidenced by decreased weight 

gain (47% gain in the benzene-treated dams vs. 69% gain in the controls [p<0.001]); benzene 

alone induced retardation of fetal and skeletal growth but did not increase the incidence of 

skeletal anomalies or skeletal, internal, or external malformations or the percentage of abnormal 

survivors. The combined developmental effects of the solvents were not additive. However, 

when each solvent was administered by inhalation in combination with orally administered ASA, 

it enhanced maternal as well as fetal toxicity. In addition, benzene and its methyl derivatives 

significantly increased the frequency of ASA-induced malformations. The induction of most 

malformations was not solvent dependent; cleft lip and palate and abnormalities of the spinal 

column occurred only in response to the combined treatment with solvents and ASA. The 

investigators concluded that under the conditions of this experiment, neither benzene nor its 

alkyl derivatives induced malformations, alone or in combination, and that the fetal toxicity 

observed with all three solvents was not additive when the solvents were combined; however, all 

three solvents potentiated the toxic effects of ASA, including the induction of malformations. 

Pregnant Swiss-Webster mice (five per exposure level per progeny age group; initial age 

8–12 weeks) were exposed via inhalation to nominal vapor concentrations of 0, 5, 10, or 20 ppm 

(0, 16, 32, or 64 mg/m3

) benzene 6 hours/day on GDs 6–15 (Keller and Snyder, 1986). Two 

fetuses/litter/sex were sacrificed on GD 16, two neonates/litter/sex were sacrificed 2 days after 

birth, and one adult/litter/sex was sacrificed 6 weeks after birth to measure hematopoietic 

progenitor cells (CFU-E, BFU-E, and GM-CFU-C) from the liver (fetuses and neonates) and 

bone marrow and spleen (adults). In addition, 10-week-old progeny from litters in the control 

and mid-exposure group were exposed for 2 weeks to 10 ppm (32 mg/m3) benzene, then 

sacrificed to measure hematopoietic progenitor cells from the bone marrow and spleen. 

There was no evidence of maternal or nonhematopoietic developmental toxicity in 

benzene-exposed mice. There was a significant increase (p<0.05) in the numbers of BFU-E 

from livers of male and female fetuses exposed to the low- and mid-exposure levels. The 

following significant changes (p<0.05) were observed with respect to CFU-E: in fetuses, there 

were increases in liver CFU-E at the low- and mid-exposure levels and decreases at the highexposure level; in male neonates, there were increases and decreases in liver CFU-E at the midexposure level and increases at the high-exposure level; and in adult mice, there were decreases 

in bone marrow CFU-E and increases in spleen CFU-E in males exposed to 10 ppm (32 mg/m3

) 

91
in utero. Liver GM-CFU-C in neonates was significantly decreased (p<0.05) at the midexposure level (males only) and increased at the high-exposure level. Mice exposed to 10 ppm 

(32 mg/m3

) benzene in utero and for 2 weeks as adults had significantly (p<0.05) decreased bone 

marrow CFU-E (males only) and splenic GM-CFU-C; mice exposed to air in utero and 10 ppm 

(32 mg/m3) benzene for 2 weeks as adults had no changes in bone marrow or splenic CFU-E, but 

they had a significant (p<0.05) decrease in splenic GM-CFU-C (females only). The authors 

concluded that benzene treatment in utero induced hematopoietic alterations in fetuses that 

persisted until at least 10 weeks after birth. 

This study could be interpreted as identifying a LOAEL of 5 ppm (16 mg/m3

) for 

developmental hematopoietic effects in mice, because statistically significant changes in BFU-E 

and CFU-E were observed in the 16-day fetuses from dams exposed to 5 ppm (16 mg/m3). 

However, the responses observed were typically biphasic in nature, showing increases in BFU-E 

and CFU-E at 5 and 10 ppm (16 and 32 mg/m3

) followed by decreases at 20 ppm (64 mg/m3). 

Only five pregnant animals were used in the study per exposure dose level, and limited numbers 

of animals (two fetuses, two neonates, or one adult/litter/sex) were used for the evaluation of 

hematotoxic effects. Also, the responses did not establish a consistent pattern in the different 

ages of progeny examined. Therefore, there is a high degree of uncertainty associated with 

determining whether the effects observed at 5 ppm (16 mg/m3) are truly adverse effects. 

Keller and Snyder (1988) further examined the effects of in utero benzene on the 

developing recognizable hematopoietic precursor cells and peripheral blood cells and Hgb 

production in Swiss-Webster mice. Three separate exposure experiments were performed in 

which pregnant mice (five per group) were exposed to benzene concentrations of 0, 5, 10, or 20 

ppm (0, 16, 32, or 64 mg/m3

) in chambers 6 hours/day on GDs 6–15. Maternal toxicity was 

evaluated on the basis of morbidity, mortality, or weight loss, but maternal peripheral blood 

count was not determined. In experiment 1, the fetuses of the benzene-exposed dams were 

assayed for signs of hematotoxicity on GD 16 (two fetuses/sex/litter). In experiment 2, the 

offspring (two neonates/sex/litter) were assayed at 2 days. In experiment 3, progeny 

(one/sex/litter) were examined at 6 weeks of age. Parameters of toxicity for the fetuses and 

2-day-old neonates included the number of live, dead, and resorbed fetuses; body weights; gross 

abnormalities; RBC, WBC, and blood cell differentials; Hgb levels; and the number of liver cells 

in the hematopoietic differentiating, proliferating pool. The responses of 6-week-old progeny 

were evaluated on the basis of peripheral WBC and RBC counts, Hgb, and smears made from 

femur bone marrow or spleen to determine the number of cells in the differentiating, 

proliferating pool. Although not stated, the experimental animals may have been the same for 

both studies (Keller and Snyder, 1986, 1988), as all the details of the exposures were identical. 

Benzene exposures monitored hourly were well within target range (Keller and Snyder, 

1988). There was no evidence of maternal toxicity at any exposure level. For the fetuses, there 

were no effects on litter size, male/female ratio, or body weight or the numbers of dead, 

resorbed, or malformed. Significant findings (p<0.05) relevant to hematopoietic toxicity in the 

three groups of progeny included the following. 

Peripheral blood. Hgb levels of exposed and control mice were similar for all groups. 

Peripheral blood cell and differential cell counts of the 16-day fetuses and differential cell counts 

of the 6-week-old progeny were comparable with those of the controls. The 2-day-old neonates 

92
exposed in utero to 5 ppm (16 mg/m3

) benzene exhibited significant increases in RBC counts and 

decreases in mean corpuscular Hgb, and the 6-week-old adults exposed in utero to 5 and 20 ppm 

(64 mg/m3

) benzene had increased RBC counts. However, these effects did not occur with a 

distinct pattern, and the investigators concluded that they probably had no toxicological 

significance. Peripheral blood cell differentials from the 2-day-old neonates revealed 

significantly decreased numbers of early nucleated red cells (statistically significant at all 

concentrations); the effect was concentration dependent. The 2-day-old neonates exposed in 

utero to 20 ppm (64 mg/m3) also had decreased numbers of late nucleated red cells and increased 

numbers of nondividing granulocytes; these effects were not concentration dependent. 

Hematopoietic organs. The 16-day fetuses from dams exposed to benzene exhibited no 

changes in hematopoietic parameters of the liver at any benzene concentration. The 2-day-old 

neonates exposed in utero to 20 ppm (64 mg/m3

) benzene had changes in the liver that were 

significant at 20 ppm (64 mg/m3

). These included increase in the number of blast cells to 3.2 

times the control value, depression in the number of late nucleated red cells (polychromatic 

normoblasts and their nucleated progeny) to 70% of the control value, an increase in the number 

of lymphocytes to 1.4 times the control value, an increase in the number of nondividing 

granulocytes to 1.9 times the control value, and an increase in the numbers of dividing 

granulocytes to 2.5 times the control value. These effects were not statistically significant at 

lower concentrations. The effects on the early red cells and granulocytes were concentration 

related; the others did not show a distinct dose response. The 6-week-old adults exposed in utero 

to 20 ppm (64 mg/m3) benzene also exhibited statistically significant changes in the cells of 

hemopoietic organs. These included fewer early nucleated red cells in the liver (to 55% of the 

control value), more splenic blast cells (to 6.5 times the control value), and increased dividing 

and nondividing splenic granulocytes (to 2.9 and 3.2 times the control value, respectively). The 

effects were not dose related and, with the exception of one cell count at 5 ppm (16 mg/m3

), were 

not statistically significant at lower concentrations. 

The authors concluded that benzene induces hematotoxicity in the offspring of mice 

exposed during pregnancy, as evidenced by reductions in the numbers of early nucleated red 

cells in the peripheral blood of 2-day-old neonates at all exposure levels and by the increase in 

the number of dividing granulocytes in the liver of 2-day-old neonates and in the spleen of 6-

week-old adults. However, the toxicological significance of these results is unclear. The only 

clearly dose-related response was the decrease in early nucleated red cells in the peripheral blood 

of 2-day-old neonates. Early nucleated red cells, however, were present in the peripheral blood 

only in very young animals, as clearly indicated by the near absence of such cells in the 6-weekold adult mice. A limited number of pregnant animals (five per group) and progeny 

(two/litter/sex) were tested. Therefore, biological significance and confidence in the data are 

questionable. Thus, a NOAEL or a LOAEL cannot be derived from the results of the Keller and 

Snyder (1988) report. 

Corti and Snyder (1996) examined the influence of gender, development, pregnancy, and 

ethanol consumption on the hematotoxicity of inhaled benzene. They exposed age-matched 

male, virgin female, and pregnant Swiss-Webster mice (number of animals in each group was 

not given, but there were 12 or 13 dams) to 10 ppm (32 mg/m3

) for 6 hours/day for 10 

consecutive days (GDs 6–15 for the pregnant females). One-half of the animals also received 

5% ethanol in the drinking water during the exposure period. On day 11, bone marrow cells 

93
from the adults and liver cells from the fetuses were assayed for the numbers of CFU-E. CFU-E 

assays were also performed on bone marrow cells isolated from 6-week-old males and females 

exposed in utero. Depressions in CFU-E numbers were seen only in males. CFU-E in adult 

males was depressed to ~70% of control values by exposure to benzene, ethanol, or ethanol plus 

benzene. The action of the two agents was neither additive nor synergistic. CFU-E from fetal 

livers was significantly decreased in males exposed to benzene or benzene plus ethanol but not 

for ethanol alone. The CFU-E value for males exposed to both benzene and ethanol, however, 

was reduced by only 5% in comparison with the control value as compared with 20% for ethanol 

alone. There were no significant differences for female fetuses. Assays of CFU-E in 6-week-old 

mice exposed to benzene in utero showed reductions of as much as 50% in comparison with 

controls for males, but there were no significant reductions for females. In fact females exposed 

to benzene and ethanol in utero showed a 40% elevation of CFU-E in comparison with 

unexposed controls. For both ethanol and benzene, only a single concentration was used in these 

experiments. Thus, it is not possible to evaluate whether the results are part of a dose-response 

relationship. 

Murray et al. (1979) conducted developmental toxicity studies in mice that also included 

the evaluation of hematologic parameters in the offspring. CF-1 mice were exposed to 500 ppm 

(1597 mg/m3

) benzene 7 hours/day on GDs 6–15 and were sacrificed on GD 18. There were no 

significant signs of maternal toxicity, and values for PCV, RBCs, Hgb, and WBCs were 

comparable with control values for the adult mice. There were no adverse effects on 

developmental parameters such as number of live fetuses or number of resorptions/litter; 

however, body weight of the mouse fetuses was slightly but significantly decreased (to 94% of 

control values; p<0.05). The investigators described a “significant” increase in the incidence of 

skeletal variations of the mouse fetuses, as evidenced by delayed ossification of sternebrae and 

skull bones and unfused occipital bones of the skull, but there was no significant increase in the 

incidence of malformations. The investigators concluded that benzene did not induce 

malformations in mice at 500 ppm (1597 mg/m3

), but it was toxic to the fetuses. This study used 

only one dose, and statistical analysis for the variations were not clearly stated; however, the 

study did demonstrate fetal toxicity at a dose that was not maternally toxic. 

Ungvary and Tatrai (1985) examined the developmental toxicity of benzene in CFLP 

mice. Groups of 11 to 15 pregnant animals inhaled 156 or 313 ppm (500 or 1000 mg/m3

) 

benzene 24 hours/day during GDs 6–15. Unexposed and air-exposed (chamber) animals served 

as controls. The mice were sacrificed on GD 18. The authors mentioned in the abstract of their 

paper that all solvents caused moderate and concentration-dependent maternal toxicity. This was 

the only information provided for maternal toxicity among the mice. Of the mouse fetuses from 

dams exposed to 156 and 313 ppm (498 and 1000 mg/m3

), 25 and 27%, respectively, exhibited 

weight retardation, and 10 and 11%, respectively, had retarded skeletal development. All of 

these effects were statistically significant (p<0.05). There were no differences between control 

and exposed groups in the incidence of minor anomalies or malformations. The investigators 

concluded that under the conditions of this study, benzene does not induce malformations, but it 

does induce fetal toxicity in mice exposed to 156 and 313 ppm (498 and 1000 mg/m3

), a 

maternally toxic dose. The mouse data suggest a LOAEL of 156 ppm (498 mg/m3). 

Murray et al. (1979) conducted developmental toxicity studies in rabbits that also 

included the evaluation of hematologic parameters in the offspring. New Zealand rabbits were 

94
exposed to 500 ppm (1597 mg/m3

) benzene 7 hours/day on GDs 6–18. The rabbits were 

sacrificed on GD 29. The rabbits exhibited no significant signs of maternal toxicity, and there 

were no adverse effects on developmental parameters such as number of live fetuses or number 

of resorptions/litter. Body weight and length of the rabbit fetuses were comparable with control 

values. Two minor skeletal variants, lumbar spurs and the proportion of fetuses with 13 ribs (the 

normal number is 12 or 13), occurred significantly less often among the exposed litters. 

Increases in the incidence of malformations were not significant. Values for PCV, RBCs, Hgb, 

and WBCs were comparable to control values for adult and fetal rabbits. The investigators 

concluded that benzene did not induce malformations in rabbits at 500 ppm (1597 mg/m3) and 

was only weakly toxic to the fetuses. This study used only one dose, and statistical analyses for 

the variations observed were not clearly stated. 

Ungvary and Tatrai (1985) also examined the developmental toxicity of benzene and 

other solvents in New Zealand rabbits. Groups of 11–15 pregnant animals inhaled 156 or 313 

ppm (500 or 1000 mg/m3

) benzene 24 hours/day on GDs 7–20. Unexposed and air-exposed 

(chamber) animals served as controls. The rabbits were sacrificed on GD 30. The maternal 

rabbits exhibited concentration-related decreases in weight gain (to 37% of control values at 313 

ppm [1000 mg/m3

]) and increases in relative liver weights (1.2 times control values at 313 ppm 

[1000 mg/m3

]); compared with control values, the effects were statistically significant at 313 

ppm (1000 mg/m3

) (p<0.05). Two dams died and six aborted at the higher concentration. The 

fetuses exhibited concentration-related decreases in body weight (to 83% of control values, 

p<0.05 at 313 ppm [1000 mg/m3

]), concentration-related increases in the percent of dead or 

resorbed fetuses (3.1 times control values, p<0.05 at 313 ppm [1000 mg/m3

]), concentrationrelated increases in skeletal retardation (not statistically significant) and minor anomalies (2.5 

times control values, p<0.05 at 313 ppm [1000 mg/m3]), and concentration-related decreases in 

the percent of malformations. The investigators concluded that under the conditions of this 

study, benzene does not cause malformations, but it induces fetal toxicity in rabbits at 313 ppm 

(1000 mg/m3), a maternally toxic concentration. The evidence supporting the LOAEL and the 

NOAEL in rabbits is weakened by a lack of experimental details in the report, the small numbers 

of pregnant rabbits employed, and the use of the fetus rather than the litter as the experimental 

unit. 

4.2.2.3. Summary of Principal Reproductive/Developmental Toxicity Effects 

As shown in the studies summarized in the previous subsections and in Tables 7 and 8, 

administration of benzene has been associated with the onset of reproductive/developmental 

effects in a wide range of experimental animals. 

The susceptibility of reproductive organs to the toxic effects of benzene was shown in 

studies on the longer-term administration of benzene in which, among other organs, the rat testes 

(Wolf et al., 1956) and the ovaries of female CD-1 mice (Ward et al., 1985; NTP, 1986) were 

reduced in weight compared with controls and histopathologically altered as a result of benzene 

administration. The potential for these changes to affect reproductive function in males is 

suggested by reduced sperm count and increased numbers of abnormal sperm when CD-1 mice 

inhaled up to 300 ppm (958 mg/m3

) benzene for 13 weeks. 

95
The experiments of Kuna et al. (1992) provided evidence that the inhalation of benzene 

by female Sprague-Dawley rats before and during pregnancy resulted in reduced body and liver 

weights in weanlings, a valid reproductive/developmental response because it occurred in the 

absence of any overt effects of the compound on maternal toxicity. In addition, a considerable 

number of experiments have addressed the potential of inhaled benzene to induce reproductive, 

developmental, or teratogenic effects in the offspring of pregnant rodents exposed to the 

compound during the period of principal organogenesis. For example, the studies by Kuna and 

Kapp (1981) of Sprague-Dawley rats point to the possibility of benzene-induced skeletal 

variations and abnormalities in the fetuses of exposed dams, although the quantitative extent of 

the effect was marginal at best. However, the weight of evidence for benzene as a 

developmental toxicant arises from the extensiveness of the overall information base. Several 

studies that employed essentially similar experimental protocols (see, e.g., Green et al., 1978) 

contribute to an overall picture of the compound as inducing reduced fetal growth and delayed 

ossification. In addition, other studies reported reductions in fetal body weights, for example, in 

Sprague-Dawley rats (Coate et al., 1984) and New Zealand white rabbits (Ungvary and Tatrai, 

1985). 

The developing hematopoietic system in mice is affected by maternal exposure to 

benzene (Keller and Snyder, 1986, 1988; Corti and Snyder, 1996), which is consistent with 

benzene-induced hematologic abnormalities in humans and adult animals. The offspring of 

dams exposed during gestation exhibited increased granulopoiesis and changes in the numbers of 

hematopoietic progenitor and precursor cells, particularly during postnatal development. 

Multigenerational animal studies on the reproductive/developmental toxicity of benzene 

were not found in the literature, an information gap that contributes to uncertainty about the 

overall context of the compound’s reproductive and developmental toxicity. As discussed in 

Sections 4.1.2.3 and 4.1.2.4, the evidence regarding reproductive and developmental effects from 

human studies is limited. The human studies suffer from major limitations, including small 

numbers of subjects, exposures to a mixture of chemicals, and poorly defined benzene exposure 

levels and durations of exposure. In many cases data on controls are lacking. 

4.2.2.4. Mechanisms of Developmental and Reproductive Toxicity 

The mechanisms for the developmental and reproductive toxicity of benzene are not well 

understood. Summarized below are a few of the suggested mechanisms that pertain specifically 

to the developmental and reproductive toxicity of benzene. 

& Following administration of benzene to pregnant rats, Pushkina et al. (1968) observed 

decreased ascorbic acid content in the whole fetus and in maternal organs as the 

concentration of benzene increased, first in the maternal liver and later in the placenta 

and fetal liver. Benzene also increased the DNA content and decreased the RNA content 

in the placenta, fetal liver, and fetal brain and it decreased DNA content in the maternal 

liver. The authors suggested that these alterations in ascorbic acid, RNA, and DNA 

content are possible mechanisms for fetal toxicity. 

& Ungvary and Donath (1984) suggested that damage to the peripheral noradrenergic fibers 

observed in their study in pregnant rats may result in a disturbed control of ovarian and 

96
uterine blood flow and steroid production and may thus be instrumental in the 

embryotoxic action of organic solvents. 

& Tatrai et al. (1980) suggested that several factors may be responsible for the 

embryotoxicity of benzene. First, because of its lipophilicity, benzene can pass the 

placental barrier and affect the embryonal cells directly. Second, phenol, a major 

metabolite of benzene shown to inhibit DNA synthesis in bone marrow in vivo, can also 

pass the placental barrier. Third, benzene can damage maternal circulation and cause 

bone marrow depression, resulting in adverse nutritional conditions for the fetus. 

Benzene oxide was suggested as the toxic metabolite of benzene, and it can be 

translocated from liver to blood and bone marrow. The enzymes responsible for its 

production appear in the rat fetus late in pregnancy. However, in the human fetus, the 

enzymes are present during weeks 9–13 of pregnancy. The authors suggested that a 

transplacental effect is more plausible than any other mechanism. 

& Reports indicating that paternal (as well as maternal) exposure to benzene is associated 

with increased risk for stillbirths and the findings of increased incidences of testicular 

lesions in benzene-exposed animals suggest that exposure of males may be important in 

the reproductive toxicity of benzene (Savitz et al., 1989; Spano et al., 1989; Ward et al., 

1985). 

4.2.3. Neurotoxicity 

4.2.3.1. Oral Exposure 

Sprague-Dawley rats given a single dose of 1870 mg/kg benzene exhibited tremors and 

tonic-clonic convulsions and died within minutes. A dose of 352 mg/kg produced slight 

nervous system depression (Cornish and Ryan, 1965). The LD50 for benzene in nonfasting rats is 

0.81 g/kg. 

Hsieh et al. (1988a) evaluated the effects of benzene on neurotransmitters of the brain. 

CD-1 adult male mice (five/group) received benzene in the drinking water for 4 weeks ad lib. 

Controls received tap water. The nominal concentrations were 0, 40, 200, and 1000 mg/L. 

However, based on water consumption, the actual daily estimated doses of benzene were 0, 8, 

40, and 180 mg/kg/day. 

The benzene concentrations used in this study did not significantly alter behavior, body 

weights, or food and water consumption. Generally, the increases in levels of monoamine 

neurotransmitters were dose related at 8 and 40 mg/kg/day, but in several cases there were no 

further increases at 180 mg/kg/day. The increases were greatest for norepinephrine in the 

hypothalamus (increased over control by approximately 38, 55, and 58% at 8, 40, and 180 

mg/kg/day, respectively) and medulla oblongata (16, 42, and 20%). For serotonin, a similar 

pattern of increase was observed in the higher association center with increased exposure to 

benzene. In the hypothalamus, the serotonin levels were 21, 86, and 93% above the control 

values. A generalized increase was also observed in the medulla oblongata (5, 25 and 19%) and 

the midbrain (8, 46, and 23%). The increases in the parent compounds were associated with 

increases in their corresponding metabolites, reflecting increased turnover of the amines. The 

97
data from this study did not establish a no-effect level of benzene in drinking water. In fact, 

although the assay has potential as a biomarker for exposure, the biological significance of these 

findings is questionable. The findings, however, suggest that neurotoxicological effects in the 

hypothalamus and the peripheral nervous system could be of particular concern for the 

developing central nervous system, resulting in permanent effects. 

4.2.3.2. Inhalation Exposure 

Ten rabbits exposed to ~ 45,000 ppm (143,760 mg/m3

) benzene exhibited light narcosis 

after 3.7 minutes of exposure, followed by tremors, chewing, excitement, and running 

movements after 5 minutes (Carpenter et al., 1944). There was a loss of pupillary reflex to 

strong light after 6.5 minutes, loss of blink reflex to tactile stimulus after 11.4 minutes, pupillary 

contraction after 12 minutes, involuntary blinking after 15.6 minutes, and death after 36.2 

minutes (Carpenter et al., 1944). 

Andersson et al. (1983) examined the effects of high concentrations of benzene on 

dopamine and noradrenaline turnover within various parts of the hypothalamus that are involved 

in neuroendocrine regulation. Turnover of the catecholamines was assessed by measuring 

changes in the degree of catecholamine stores following tyrosine hydroxylase inhibition. 

Sprague-Dawley rats exposed to 1500 ppm (4792 mg/m3

) benzene 6 hours/day for 3 days were 

divided into two groups of four or six animals. In the first group (–inhibitor), rats that served as 

controls were sacrificed immediately following exposure. In the second group (+inhibitor), the 

injection of tyrosine hydroxylase inhibitor immediately followed exposure, and the animals were 

sacrificed 2 hours later. The extent of depletion of the catecholamines was determined by 

calculating the percentage depletion in the test group, based on the levels of catecholamine 

present in the control group at the time of the injection of the hydroxylase inhibitor. 

Quantitative microfluorimetry demonstrated that benzene induced statistically significant 

alterations in catecholamine content and turnover in various sections of the hypothalamus. 

Benzene (–inhibitor) produced increases in the catecholamine fluorescence in the median and 

lateral palisade zones of the median eminence (p<0.05) and within the posterior periventricular 

hypothalamic region (p<0.05). Benzene (+inhibitor) enhanced the disappearance of 

catecholamine in the median palisade zone (p<0.002), within the posterior periventricular 

hypothalamic region (p<0.05), within the parvocellular part of the paraventricular hypothalamic 

nucleus (p<0.01), and within the dorsomedial hypothalamic nucleus (p<0.01). 

The investigators concluded that benzene produced a pattern of discrete changes in 

noradrenalin and dopamine turnover in certain areas of the hypothalamus. The study included 

only one concentration of benzene and a small number of animals. No LOAEL or NOAEL was 

established. Tyrosine hydroxylase is the key enzyme for biosynthesis of catecholamine, and the 

hypothalamus is one of the major association centers in the central nervous system. It is clear 

that benzene will affect the functions of all these centers at an inhalation dose of 500 ppm (1597 

mg/m3

), as demonstrated in the changes in dopamine levels of various areas, probably including 

centers controlling respiration, hunger, and thirst. 

Ungvary and Donath (1984) evaluated the effects of benzene on the noradrenergic 

innervation of reproductive organs. CFY rats were exposed to 465 (1500 mg/m3

) benzene 8 

98
hours/day on GDs 8–10 ppm. The abundance of fluorescent noradrenergic fibers normally found 

in the ovaries and uterus of the pregnant rats decreased, whereas background fluorescence 

increased (interpreted by the investigators to indicate an increased release of noradrenalin). To 

confirm that the effect was selective to nerve fiber, the investigators injected benzene into the 

anterior chamber of the eye. The benzene-induced damage to the sympathetic nerve-plexus of 

the iris was similar to that in the uterus and ovary. The density of fluorescent fibers was 

decreased (dose related), and the iris exhibited substantial hyperemia and increased background 

fluorescence 72 hours after injection. The investigators concluded that benzene has a selective 

and differential toxic effect on postganglionic neurons, with potential embryotoxicity. 

Dempster et al. (1984) examined the temporal relationship between the behavioral and 

hematologic effects of inhaled benzene. Mice (30–45/group) were exposed to benzene 

concentrations of 100–3000 ppm (319–9584 mg/m3

) 6 hours/day for the number of days 

necessary to achieve a concentration × time product of 3000 ppm-days. The controls were agematched and exposed to air. The behavioral parameters evaluated included milk licking (an 

observation of the mice licking a spout protruding from a wall in order to obtain milk), hindlimb 

grip strength (a mouse was held by the tail and steadily pulled backward through a trough until 

both hind paws grasped a wire triangle and then it was pulled until its grip was broken), and 

home cage food and water intake (measured by weighing water bottles and feeders daily). 

Lymphocyte numbers were reduced to 68% of control values after five exposures to 100 ppm 

(319 mg/m3

) benzene, to 50% after two exposures to 300 ppm (958 mg/m3), and to 50% after one 

exposure to 1000 and 3000 ppm (3194 and 9584 mg/m3

). Maximal depression in the lymphocyte 

counts occurred after 10 days of exposure to 300 ppm (958 mg/m3

). The lymphocyte counts 

remained depressed throughout each exposure regimen. 

RBC counts were not depressed as rapidly or to the same extent as were the lymphocytes. 

At the end of exposure to 300 ppm (958 mg/m3

), for example, the RBC counts were reduced to 

70% of control values. Increased milk licking, the most sensitive of the behavioral parameters, 

was statistically significant following 1 or 2 days of exposure to 100 ppm (319 mg/m3

) and after 

4 or 5 days of exposure to 300 ppm (958 mg/m3

). The maximal increase in this behavior pattern 

occurred after 7–8 days of exposure to 300 ppm (958 mg/m3), following the same time course as 

the hematologic effects. Short-term exposures to high concentrations did not increase milk 

licking, but they did increase food consumption. Thus the increased milk licking at 100 ppm 

(319 mg/m3

) was apparently not due to hunger. Exposure to 1000 ppm (3195 mg/m3) for one 

exposure reduced hindlimb grip strength. These effects disappeared following termination of 

exposure. 

In a study designed to reflect occupational exposure, male CD-1 and C57BL/6 mice were 

exposed to 300 or 900 ppm (958 or 2875 mg/m3

) benzene 6 hours/day for 5 days followed by 2 

weeks of no exposure, after which the exposure regimen was repeated for an unspecified amount 

of time (Evans et al., 1981). Seven categories of behavioral activities were monitored in exposed 

and control animals: stereotypic behavior, sleeping, resting, grooming, eating, locomotion, and 

fighting. Only minimal and insignificant differences were observed between the two strains of 

mice. Increased behavioral activity in both strains was observed after exposure to benzene. 

Mice exposed to 300 ppm (958 mg/m3) benzene had a greater increase than those exposed to 900 

ppm (2875 mg/m3

), probably because of narcosis-like effects induced at the higher exposure 

99
level. It is not known whether benzene induces behavioral changes by direct action on the 

central nervous system. 

Frantik et al. (1994) found that male albino SPF rats from a Wistar-derived strain 

exposed to benzene for 4 hours in glass chambers (dose not specified) exhibited depression of 

evoked electrical activity in the brain; the authors calculated the 30% effect level (depressed 

activity) as 929 ppm (2968 mg/m3). When female H strain mice were exposed to benzene for 2 

hours, the 30% effect level for depression of evoked electrical activity in the brain was 856 ppm 

(2735 mg/m3

). 

Adult male Kunming mice (five/group) were exposed to 0, 0.78, 3.13, or 12.52 ppm (0, 

2.5, 10, 40 mg/m3

) benzene 2 hours/day, 6 days/week for 30 days (Li et al., 1992). Exposures 

were conducted under static conditions in 300 m3 plexiglass chambers. Benzene was monitored 

by gas chromatography every 30 minutes for 3 days at the beginning of the experiment, but 

apparently it was not monitored for the rest of the experiment. The animals were monitored for 

Y maze performance (rapid response); locomotor activity; forelimb grip strength; acetyl 

cholinesterase (AChE) activity in the blood and brain; brain, liver, spleen, and kidney weights; 

and bone marrow cellularity. Statistical significance was evaluated by t value for forelimb grip 

strength and locomotor activity and by U value for rapid response frequency. 

Significantly increased grip strength was observed at 0.78 ppm (2.5 mg/m3

), but at higher 

doses, grip strength was significantly decreased. The frequency of rapid response (running a Y 

maze in less than 3 seconds) followed a pattern similar to forelimb grip strength. The frequency 

increased significantly (p<0.05) from 33.7% in controls to 43.8% in the 0.78 ppm group but 

declined significantly to 29.4% and 25.5% in the 3.13 and 12.52 ppm (10 and 40 mg/m3

) groups, 

respectively. No statistically significant differences were observed in AChE activity in either 

blood or brain or in locomotor activity at any dose level. Relative liver weight was significantly 

increased and relative spleen weight was significantly (p<0.05) decreased in the high-dose 

group. No statistical analysis of the bone marrow histologic investigation was presented. There 

were no apparent responses in the 0.78 or 3.13 ppm (2.5 or 10 mg/m3

) groups. In the high-dose 

12.52 ppm (40 mg/m3) group, however, there were reductions of 91% in myeloblasts, 64% in 

premyeloblasts, 77% in metamyelocytes, 100% in reticulum, and 100% in erythroblasts. 

The neurological effects reported in this study—forelimb grip strength and frequency of 

rapid response—are unique and interesting. However, several limitations in this study prevent 

its use in establishing an RfC. The number of animals used in each group (five) was low, 

exposures were performed under static conditions, and benzene concentrations were monitored 

on only the first 3 of 30 exposure days. The very large decreases in several blood parameters are 

in contrast with the findings of most other studies, which found minimal or no response in bone 

marrow parameters at similar exposure concentrations. Most other studies on the hematotoxicity 

of inhaled benzene used exposure durations of 6 hours/day, in contrast to the 2 hours/day in this 

study. Thus, the dramatic effects on bone marrow parameters in this study suggest that actual 

benzene exposure may have been higher than reported. 

100
4.2.3.3. Summary of Neurotoxic Effects 

The database for establishing threshold dose levels for neurotoxicity effects is limited. 

Neurological effects of benzene have been observed in animals and humans, and no clear 

NOAEL could be identified in any species. No complete neurological testing has been 

conducted in animals or humans. The exposure levels used in the neurotoxicity studies 

summarized here were typically high and the exposure durations were short. The longest 

exposure period was 4 weeks (Hsieh et al., 1988a). It should be noted that neurological concerns 

have not been raised in occupational studies where exposure was at lower levels. Kahn and 

Muzyka (1973) reported neurological complaints in workers exposed to 6–16 ppm 

(20–52mg/m3), but no objective evaluation was conducted, exposure was poorly characterized, 

and there was confounding exposure to low levels of toluene (1.6 ppm [5 mg/m3]). Other human 

neurotoxicity studies reported much higher exposure levels. No complete functional 

observational battery has been conducted in animals. Dempster et al. (1984) reported reduced 

milk licking in mice exposed to 100 ppm (319 mg/m3

 for 1 or 2 days, but grip strength was not 

affected until exposure reached 1000 ppm (3195 mg/m3). Li et al. (1992) reported effects on grip 

strength, locomotor activity, and Y-maze performance (rapid response) at 0.78 ppm (2.5 mg/m3

), 

but poor exposure characterization and marked inconsistencies between this and other studies’ 

hematologic effects mean that this value is unreliable. 

4.2.4. Immunotoxicity 

4.2.4.1. Oral Exposure 

Male Charles River CD-1 mice (five/group, 6–7 weeks of age) were exposed to 0, 31, 

166, or 790 mg/L (0, 8, 40, or 180 mg/kg/day) benzene in drinking water for 28 days (Hsieh et 

al., 1988b). The treatment had no adverse effects with respect to mortality, clinical signs, body 

weight change, liver weight, or gross necropsy. A dose-related decrease in relative spleen 

weight was observed, significant (p<0.05) at the high-exposure level. In one test, spleen 

cellularity was reported to be significantly decreased at all exposure levels, and in a separate test 

only at the high-exposure level. Although relative thymus weights were decreased at all 

exposure levels, the values were not statistically significantly different from control values. 

Dose-related hematologic effects (erythrocytopenia, leukocytopenia, lymphocytopenia, increased 

MCV) were observed at all exposure levels. The authors indicated that the increased MCV and 

decreased HCT and numbers of RBCs were indicative of severe macrocytic anemia. 

Biphasic responses were observed in immunological tests, including mitogen-stimulated 

(lipopolysaccharide [LPS], pokeweed mitogen, concanavalin A [Con A], phytohemaglutinin 

[PHA]) splenic lymphocyte proliferation; mixed splenic lymphocyte culture response to 

allogenic yeast artificial chromosome [YAC]-1 cells; cytotoxic splenic T lymphocyte response to 

allogenic YAC-1 cells with a significantly increased response at the low-exposure level; and 

significantly decreased responses at the mid- and/or high-exposure level. Using several methods 

to determine primary antibody response to SRBC, significantly decreased responsiveness was 

observed at the mid- and/or high-exposure levels. This response was either significantly 

increased or not different from controls in mice at the low-exposure level. This study identified 

a LOAEL of 8 mg/kg/day (the lowest dose tested) for hematologic and immunological effects in 

male mice exposed to benzene in drinking water for 30 days. No NOAEL was established. 

101
In a subchronic study, groups of male C57BL/6 mice were exposed via drinking water to 

0, 152, or 853 mg/L benzene for 7–28 days. Using estimated daily water intakes, the authors 

calculated benzene dosages of 27 and 154 mg/kg/day (Fan, 1992). Five mice per group were 

sacrificed after 7, 14, 21, or 28 days of exposure. An unspecified number of mice were exposed 

to 152 mg/L for 28 days and sacrificed 7, 14, or 21 days after the last dosage. The focus of this 

study was to determine the toxicity of benzene on natural killer (NK) cells involved in 

nonspecific host resistance and on IL-2, which is the primary growth factor of T cells, a growth 

factor for B cells and NK cells, and is involved in the regulation of granulocyte and eosinophil 

production that occurs in response to NK cell activity and IL-2 production. No overt signs of 

toxicity were observed in the benzene-exposed mice. Significant decreases in the number of 

spleen cells were observed in both groups of benzene-exposed mice. This effect was observed 

after 21 days of exposure in the 152 mg/kg/day group and after 14 days in the 852 mg/kg/day 

group. After 21 days, a significant increase in splenic NK cell activity was observed in both 

groups; however, after 28 days, the activity was not significantly different from that of controls. 

Splenic IL-2 production was significantly depressed after 28 days in both groups. Spleen cell 

numbers and IL-2 production were also depressed in the mice exposed to 152 mg/kg/day for 28 

days and sacrificed 7 and 14 days (IL-2 levels only) after the end of the exposure. This study 

identified a LOAEL of 152 mg/kg/day (the lowest dose tested) for effects on the immune system 

in male mice. A NOAEL was not identified. 

Female B6C3F1 mice (12/group, 6–7 weeks of age) were exposed to benzene in drinking 

water (containing emulphor to increase solubility of benzene) at levels of 0, 50, 1000, or 2000 

mg/L (0, 12, 195, or 350 mg/kg/day, as calculated by the authors) for 30 days (White et al., 

1984). Body weight was significantly decreased at the high-exposure level (p<0.05). A doserelated decrease in absolute and relative spleen weight was observed (p<0.01). In one test, 

spleen cellularity was reported to be significantly decreased at all exposure levels (p<0.05), and 

in a separate test at only the mid- and high-exposure levels. Dose-related leukopenia and 

lymphocytopenia were observed (p<0.05). A dose-related decrease in eosinophils was observed 

(p<0.01). At the high-exposure level, significant decreases in levels of erythrocytes and Hgb 

were observed (p<0.05). No exposure-related effects were observed for levels of blood urea 

nitrogen, serum creatinine, serum glutamic oxaloacetic transaminase, or serum glutamic pyruvic 

transaminase, indicators of renal and hepatic damage. 

Dose-related changes were observed in immunological tests on spleen cells and in assays 

of bone marrow (p<0.05); decreases were observed with respect to IgM antibody forming 

cells/spleen in response to SRBC, lymphocyte proliferation response to the T cell mitogen Con 

A and the B cell mitogen LPS, number of T lymphocytes, and femoral GM-CFU; and an increase 

was observed in bone marrow cell DNA synthesis. These effects were not significant at 12 

mg/kg/day but were dose related (p<0.05). Of all the immunological indices tested, only one 

endpoint (stimulation index for lymphocyte proliferation of spleen cells in response to medium 

containing 0.5 g/mL Con A) was significantly decreased at 12 mg/kg/day (p<0.05). The 

number of B lymphocytes was not affected, but the investigators commented that the number of 

B lymphocytes in the controls was lower than for historical controls for their laboratory. This 

study identifies a NOAEL of 12 mg/kg/day and a LOAEL of 195 mg/kg/day for hematologic 

effects in mice exposed to benzene in drinking water for 30 days, and a LOAEL of 12 mg/kg/day 

for immunological effects. 

102
4.2.4.2. Inhalation Exposure 

Male C57BL/6J mice (7–8/group) were exposed to benzene at a concentration of 0, 10.2, 

31, 100, or 301 ppm (0, 32.6, 99, 319, or 962 mg/m3

) in whole-body dynamic inhalation 

chambers 6 hours/day for 6 days (Rozen et al., 1984). Mice were bled within 30–90 minutes of 

the last exposure to determine peripheral blood counts. Five animals with blood counts closest 

to the group mean were selected for lymphocyte assays. Single-cell suspensions of bone marrow 

and spleen cells from these animals were used for T and B cell enumeration and mitogeninduced proliferative assays. Statistical significance (p<0.05) was determined by one-way 

ANOVA. 

Lymphocyte counts were depressed at all exposure levels. Lymphocyte counts were 

significantly reduced to approximately 65% of controls in the 10.2 ppm (32.6 mg/m3

) exposure 

group and to 35% of controls at 100 or 300 ppm (319 and 962 mg/m3

). Erythrocyte counts were 

significantly stimulated (115% of control levels) at 10.2 ppm (32.6 mg/m3

) and were depressed 

only at 100 and 300 ppm (319 and 962 mg/m3). At exposures of 10.2 ppm (32.6 mg/m3), the 

frequency of femoral B lymphocyte colony-forming cells was reduced to approximately 30% of 

the control value. In contrast, the number of femoral B lymphocytes was not significantly 

reduced in the low-exposure group, but the number was reduced to less than 10% of the control 

value at 100 or 300 ppm (319 and 962 mg/m3

). Similarly, splenic PHA-induced blastogenesis 

was significantly depressed at 31 ppm (99 mg/m3

) without a concomitant depression in numbers 

of T lymphocytes. The numbers of T lymphocytes were reduced to less than 50% of controls in 

the 100 and 300 ppm (319 and 962 mg/m3

) exposure groups. These results demonstrate that 

short-term exposure to inhaled benzene even at low exposure concentrations can cause 

reductions in immune-associated processes. This study identifies a LOAEL for depression of 

lymphocytes of 10.2 ppm (32.6 mg/m3), the lowest dose tested. 

Rosenthal and Snyder (1985) investigated the effects of exposure to benzene at 0, 10, 30, 

100 (320 mg/m3) or 300 ppm (960 mg/m3) on the immune response of male C57BL/6J mice 

(5–7/group) to challenge with the facultative intracellular pathogen Listeria monocytogenes. All 

mice (5–7/group) were exposed to benzene for 5 days (6 hours/day) before infection with L. 

monocytogenes. At this point benzene exposure was stopped for half of the groups, and the other 

half continued to be exposed for 7 days after infection, for a total of 12 days’ exposure. 

Bacterial proliferation in the spleen was measured at 1, 4, and 7 days after infection as an index 

of the resistance of the mice to infection. Body and spleen weights were determined, T and B 

lymphocytes were enumerated, spleen cellularity was determined, and spleen 

monocyte/macrophage, polymorphs, lymphocytes, and nucleated RBCs were scored. 

None of the benzene exposure treatments affected L. monocytogenes counts in the spleen 

1 day after infection. Preexposure to benzene at 300 ppm resulted in a sevenfold increase in 

spleen bacterial counts at 4 days after infection, but lower concentration had no significant 

effect. However, with continued benzene exposure after infection, concentration of 30 ppm (96 

mg/m3

) or more resulted in dose-dependent increases in spleen bacterial counts at 4 days after 

infection. By 7 days after exposure, spleen bacterial counts had returned to control levels in all 

treatments. The authors suggested that benzene exposure caused a delay in the cell-mediated 

immune response, as there was a temporary increase in spleen bacterial counts. Both T and B 

103
lymphocytes were particularly sensitive to benzene exposure. Lymphocyte counts were 

depressed at benzene exposure concentrations of 30 ppm, and counts did not return to control 

levels even after cessation of benzene exposure. This study identified a LOAEL of 30 

(96 mg/m3

) ppm and a NOAEL of 10 (32 mg/m3) ppm for effects on the immune system. 

4.2.4.3. Summary of Immunotoxic Effects 

As detailed above, dose-related adverse effects on spleen weight and cellularity and on 

various measures of immune function have been observed following both oral and inhalation 

exposures to benzene. Dose-related decreases in spleen weight and cellularity were consistently 

observed following oral benzene exposure (Hsieh et al., 1988b: Fan, 1992; White et al., 1984). 

Dose-related adverse effects on several measures of immune function have been observed, 

including lymphocyte count, mitogen-stimulated lymphocyte proliferation, primary antibody 

response to SRBCs, IL-2 production, femoral B lymphocyte colony-forming cells, number of T 

lymphocytes, and proliferation of L. monocytogenes bacteria in the spleen (Hsieh et al., 1988b; 

Fan, 1992; White et al., 1984; Rozen et al., 1984). The results indicate that exposure to benzene, 

whether oral or inhaled, adversely affects the immune response. 

4.3. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS AND 

MODE OF ACTION 

Benzene exposure results in adverse noncancer effects by all routes of administration. 

Hematotoxicity and immunotoxicity have been consistently reported to be the most sensitive 

endpoints of noncancer toxicity in limited studies in humans and experimental animals, and these 

effects have been the subject of several reviews (Aksoy, 1989; Goldstein, 1988; Snyder et al., 

1993b; Ross, 1996). The bone marrow is the target organ for the expression of benzene 

hematotoxicity and immunotoxicity. 

Chronic exposure to benzene results in progressive deterioration in hematopoietic 

function. Anemia, leukopenia, lymphocytopenia, thrombocytopenia, pancytopenia, and aplastic 

anemia have been reported after chronic benzene exposure (see detailed discussions and 

summaries in Sections 4.1.2.1 and 4.2.1 and Tables 1 and 6). In contrast to these blood 

cellularity depression effects, benzene is also known to induce bone marrow hyperplasia. AML 

has been frequently observed in studies of human cohorts exposed to benzene, and there is 

evidence linking benzene exposure to several other forms of leukemia (U.S. EPA, 1998a). 

Whether the hematotoxic/immunotoxic effects of benzene and its carcinogenic effects are caused 

by a common mechanism is not yet known, due in part to the fact that although the bone marrow 

depressive effects of benzene in humans can be readily duplicated in several experimental animal 

model systems, a suitable experimental animal system for the induction of leukemia has not yet 

been developed. In addition, the hematotoxicity/immunotoxicity of benzene leads to significant 

health effects apart from potential induction of leukemia. 

Although the decreased ALC and leukemias observed from benzene exposure both result 

from bone marrow toxicity, they do not necessarily result from the same mechanisms, and 

decreased ALC may not be a necessary precursor for leukemia. For example, decreased ALC 

may be the result of cytotoxicity independent of the interaction of benzene metabolites with 

DNA and/or DNA-associated proteins and tumor formation. Thus, integration of the cancer and 

104
noncancer health assessments of benzene is not considered possible given the current state of 

knowledge regarding mode of action. 

Leukocytopenia has been consistently shown to be a more sensitive indicator of benzene 

toxicity than anemia in experimental animal systems, and lymphocytopenia has been shown to 

be an even more sensitive indicator of benzene toxicity than is overall leukocytopenia (Snyder et 

al., 1980; Ward et al., 1985; Baarson et al., 1984). Rothman et al. (1996a) also found that a 

decrease in absolute lymphocyte count was the most sensitive indicator of benzene exposure in a 

group of workers. Similarly, Ward et al. (1996) demonstrated a strong relationship between 

benzene exposure and decreased WBC counts in a rubber worker cohort, but no significant 

relationship with RBC counts was found. A common observation in experimental animal and 

human studies on benzene and lymphocytopenia—except for the occupational study of Collins et 

al. (1991)—has been the absence of a clear threshold. 

Human and experimental animal studies indicate that benzene affects both humoral and 

cellular immunity. Dose-related adverse effects on spleen weight and cellularity and various 

parameters of immune function have been observed following both oral and inhalation exposure 

to benzene. 

Benzene toxicity to humans exposed in the workplace has been characterized as having 

either early hematotoxicity or, with prolonged exposure to high doses, irreversible bone marrow 

damage. Studies of worker populations in factories in which benzene was used as a solvent 

(Snyder and Kocsis, 1975) showed a range of hematotoxic effects, including anemia, leukopenia, 

and thrombocytopenia. In some cases, more than one cell type was decreased. A decrease in the 

levels of all the classes of blood cell types in circulation is termed pancytopenia and is usually 

associated with irreversible bone marrow aplasia. Aplastic anemia is fatal in most cases. In 

those who survive aplasia, the marrow appears to be dysplastic. Myelodysplastic syndrome, 

which has been called preleukemia, is likely an early stage of AML. 

The evidence is strong that benzene metabolism plays a critical role in toxicity (Snyder 

and Hedli, 1996). Hepatic metabolism plays an important role in toxicity (Andrews et al., 1977). 

In addition to hepatic metabolism, it appears that secondary metabolism of benzene metabolites 

in bone marrow contributes to toxicity (Irons et al., 1980; Schlosser and Kalf, 1989). Thus, 

elucidation of the metabolic pathway for benzene biotransformation is essential for a full 

understanding of the mechanism of toxicity. 

There are alternative routes by which the first step of benzene metabolism, namely, 

phenol formation, can occur. CYP2E1, and perhaps other cytochromes P450, can generate H2O2 

when acting as oxidases of nicotinamide adenine dinucleotide phosphate (NADPH). The 

hydroxyl radical formed from H2O2 can hydroxylate benzene to yield phenol. An alternative 

mechanism for phenol formation reflects on the fate of the benzene oxide-oxepin system (see 

Figure 1). When benzene oxide is the first product, it can rearrange nonenzymatically to form 

phenol. Alternatively, benzene oxide can be further metabolized by epoxide hydrolase to yield 

1,2-benzene dihydrodiol, which can in turn be oxidized via dihydrodiol dehydrogenase to form 

catechol. The reaction of benzene oxide with GSH catalyzed by GSHS transferase leads to the 

formation of the premercapturic acid. It is likely that benzene oxide or its oxepin (intermediate) 

are precursors to ring opening (Witz et al., 1996). Phenol can be further hydroxylated to form 

105
hydroquinone or catechol. In theory, 1,2,4-benzenetriol may be formed by the hydroxylation of 

either hydroquinone or catechol, but Inoue et al. (1989) suggested that catechol is not a precursor 

of 1,2,4-benzenetriol in humans. 

Any of the phenolic metabolites may be conjugated with either sulfate or glucuronide. In 

addition to L-phenylmercapturic acid reported by Parke and Williams (1953), other 

mercapturates include 6-N-acetylcysteinyl-S-2,3-cyclohexadienol (Sabourin et al., 1988a, b) and 

2,5-dihydroxy-phenyl-mercapturic acid (Nerland and Pierce, 1990). The urine also contains two 

ring-opening products, MA (Drummond and Finar, 1938) and 6-hydroxy-trans, trans-2,4-

hexadienoic acid (Kline et al., 1993), and the residue of a covalently bound DNA adduct, i.e., 

N7

-phenylguanine (Norpoth et al., 1988). 

The production of benzene metabolites, largely in the liver, is followed by their transport 

to the bone marrow and other organs. There are many possible causes of bone marrow toxicity. 

Pfeifer and Irons (1983) suggested that covalent binding of hydroquinone to spindle fiber protein 

could explain inhibition of cell replication by benzene. Damage to DNA could result in bone 

marrow depression, leading to aplastic anemia, which in survivors leads to marrow dysplasia and 

ultimately to AML (Snyder and Kalf, 1994). Benzene metabolites could cause damage to DNA 

by two possible mechanisms. One pathway focuses on the metabolic activation of benzene to 

species that covalently bind to DNA to produce mutagenic events that are expressed as leukemia. 

The second mechanism involves the production of metabolites that cause oxidative stress, 

subsequent oxidative damage to DNA, and a mutagenic effect that has the same consequences. 

Within the bone marrow, both hematopoietic progenitor cells and bone marrow stromal 

cells are potential targets of benzene toxicity (Aksoy, 1988; Gaido and Wierda, 1985). 

Progenitor cells are thought to be the cells of origin for leukemias (Greaves, 1993) and are 

attractive as potential targets of toxins such as benzene, whose toxic effects are not restricted to a 

single hematopoietic lineage. Stromal cells are critical in the regulation of normal hemopoiesis 

(Dorshkind, 1990) and have been considered as important targets of benzene toxicity (Gaido and 

Wierda, 1985). Stromal cells are intimately associated with developing blood cells and regulate 

hemopoiesis via direct cell-to-cell interactions, the production of extracellular matrix 

components, and the secretion of soluble mediators such as cytokines and eicosanoids (Billips 

et al., 1991). 

Although metabolism is central to benzene toxicity, studies of the metabolic capability of 

human bone marrow are scarce, particularly in human stromal and CD34+

 cells. 

Benzene hematotoxicity is a complex process that most likely involves interaction among 

several metabolites (Billips et al., 1991). Possible intermediates considered at present to be 

important in benzene hematotoxicity are polyhydroxylated benzene metabolites, for example, 

hydroquinone and 1,2,4-benzenetriol; their quinone oxidation products, for example, pbenzoquinone, formed via oxidation of hydroquinone, and semiquinone free radical 

intermediates formed during the oxidation of polyhydroxylated metabolites to quinones; reactive 

oxygen species formed during the oxidation of the polyhydroxylated metabolites; and ringopened aldehydic benzene metabolites, for example, MUC and 6-hydroxy-trans, trans-2,4-

hexadienal. 

106
MUC and its metabolite 6-hydroxy-trans, trans-2,4-hexadienal are reactive ring-opened 

hematotoxic compounds. They exhibit a host of biological activities that could potentially be 

important in their mechanisms of toxicity in relation to benzene. 

Benzene hematotoxicity occurs when its hepatic metabolites (Sammett et al., 

1979)—phenol, catechol, and hydroquinone—are transported to the bone marrow (Greenlee et 

al., 1981) and further oxidized in a peroxidase-mediated (Smith et al., 1989) reaction to 

biologically reactive intermediates that can potentially affect hematopoiesis. 

The ability to alter cytokine-dependent growth and differentiation in hematopoietic 

progenitor cells appears to be a property of many agents, with leukemogenic potential for 

humans (Irons and Stillman, 1993). Several studies have reported significant effects of benzene 

on hematopoietic stem and progenitor cells (Cronkite et al., 1989). In one study (Seidel et al., 

1989), a dose-dependent depression of all stem cell compartments was observed in BDF1 mice 

exposed for 16 weeks to airborne concentrations of benzene as high as 99 ppm (317 mg/m3

) for 6 

hours/day, 5 days/week. The results of this study indicated that the GM-CFU was much less 

sensitive than the erythroid CFUs at higher doses of benzene. 

Another study (Dempster and Snyder, 1990) reported that short-term exposure of mice to 

benzene induced a shift toward granulocytic differentiation, a growth advantage for granulocytic 

progenitor cells in the bone marrow and spleen, and to a resultant increase in the total number of 

granulocytes. These results suggest that benzene, or hydroquinone, is acting on the myeloid 

stem or progenitor cells. The administration of benzene, or hydroquinone, specifically 

stimulates granulopoiesis in mice and induces granulocytic differentiation in myeloblasts of the 

human promyelocytic leukemia cell line, HL-60, as well as the normal murine IL-3 dependent 

myeloblastic cell line, 32D.3 (G). Benzene and hydroquinone do this by replacing the 

requirement for granylocyte colony-stimulating factor and leukotriene D4, respectively, for 

induction of differentiation. 

In conclusion, benzene-induced mechanisms of hematotoxicity are poorly understood. 

Most proposed mechanisms of benzene-induced effects such as cytotoxicity, apoptosis, 

mutagenesis, and cell replication can be demonstrated in vitro at relatively high concentrations of 

benzene metabolities (Snyder et al., 1993b). Benzene effects in vitro do not necessarily correlate 

with benzene-induced hematotoxicity following in vivo exposures to low concentrations, as in 

vitro assays of hematopoiesis lack the complex hematopoietic cell interrelationships, fine-tuned 

regulation, and compensation mechanisms that are essential in normal hematopoiesis. The most 

important aspects of benzene-induced hematotoxicity are the possible toxic effects caused by 

low concentrations of benzene. Concurrent assessment of toxicity elicited by high 

concentrations of benzene (100 [320 mg/m3

] and 200 [640 mg/m3] ppm) established a reference 

for comparison. 

In common with many other organic solvents, benzene has been shown to produce 

neurotoxic effects in test animals and humans after short-term exposures to relatively high 

concentrations. The neurotoxicity of benzene, however, has not been extensively studied, and no 

systematic studies of the neurotoxic effects of long-term exposure were located. Benzene 

produces generalized symptoms such as dizziness, headache, and vertigo at levels of 250–3000 

ppm (799–9584 mg/m3

) (Brief et al., 1980), leading to drowsiness, tremor, delirium, and loss of 

107
consciousness at 700–3000 ppm (2236–9584 mg/m3

). These neurological symptoms are 

reversible upon removal of the subject from exposure. Kahn and Muzyka (1973) reported that 

workers exposed to benzene for 2–9 years at 6–15.6 ppm (20–50 mg/m3

) complained of frequent 

headaches, became tired easily, had difficulties sleeping, and complained of memory loss. The 

limitations of this study were that the workers were exposed to both benzene and toluene and the 

dose and duration of exposure were unknown. 

Few studies in test animals have examined the neurotoxicity of benzene, and those that 

have used short-term exposures, usually to concentrations that have been shown to induce 

significant hematotoxicity. One exception is the study by Li et al. (1992), which observed 

biphasic responses in forelimb grip strength and frequency of rapid response in running a Y 

maze following inhalation exposure to 0, 0.78, 3.13, or 12.52 ppm (0, 2.5, 10, or 40 mg/m3

) 

benzene 2 hours/day, 6 days per week for 30 days. Both responses increased at the lower 

concentration and declined at the intermediate and high concentrations. There were, however, a 

number of uncertainties in the experimental protocol that need to be resolved before these 

observations can be used in establishing NOAEL or LOAEL values. 

No complete neurological testing has been conducted in animals and humans. The 

toxicological significance of these responses is also not clear. A detailed discussion of 

neurotoxicity studies is presented in Section 4.2.3. 

There is some evidence of reproductive and developmental benzene toxicity from human 

epidemiology studies, but the data did not provide conclusive evidence of a link between 

exposure and effects. In most cases, there was exposure to other chemicals as well, and the 

quantitative data were not sufficient to determine a NOAEL or a LOAEL. Some test animal 

studies provide limited evidence that benzene affects reproductive organs; however, these effects 

were limited to high-exposure concentrations that exceeded the maximum tolerated dose. 

Fertility studies that have shown adverse effects on the number of live fetuses used benzene 

concentrations that caused severe maternal toxicity, as indicated by large reductions in body 

weight gain. Studies that used lower benzene concentrations showed no reduction in fertility 

(Coate et al., 1984; Green et al., 1978; Kuna et al., 1992; Murray et al., 1979). 

Results of inhalation studies conducted in test animals are fairly consistent across species 

and demonstrate that benzene is fetotoxic and causes decreased fetal weight and/or minor 

skeletal variants at concentrations of greater than 47 ppm (150 mg/m3

) (Coate et al., 1984; Green 

et al., 1978; Kuna and Kapp, 1981; Murray et al., 1979). Exposure of mice to benzene in utero 

during development has been shown to cause changes in the hematogenic progenitor cells in 

fetuses, 2-day-old neonates, and 6-week-old adults (Keller and Snyder, 1986, 1988). However, 

the biological significance of these effects is questionable because of the experimental design 

limitations (see detailed discussion of studies in Section 4.2.2). 

Although benzene exposure has been shown to result in structural and numerical 

chromosomal abberations in human lymphocytes, the quantitative relationship between 

measured benzene exposures and clastogenic effects in humans is unknown. Associations 

between benzene exposure and genotoxic effects in humans under occupational conditions have 

been demonstrated. In animal studies benzene has been shown to induce cytogenic effects, 

including chromosome and chromatid aberrations, sister chromatid exchanges, and micronuclei 

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in vivo and in vitro. Mutagenicity of benzene metabolites is well established. However, 

benzene itself is not mutagenic in bacterial and animal systems (for a detailed discussion on 

benzene genotoxicity, see Section 3.3.6.2 and U.S. EPA, 1998a). 

4.4. SUSCEPTIBLE POPULATIONS 

4.4.1. Childhood Susceptibility 

This section reviews information regarding whether infants and children may be more 

susceptible than adults to the hematotoxic effects of benzene. Developmental toxicity was 

described in Section 4.2.2. For example, inhalation exposure of benzene in pregnant rats on GDs 

6–15 resulted in fetotoxicity at 100 ppm (319 mg/m3

). The study established a LOAEL of 100 

ppm (319 mg/m3) and a NOAEL of 40 ppm (128 mg/m3) for the fetal toxicity of benzene. 

Benzene could be a potential risk factor for the development of childhood leukemia (see reviews, 

OEHHA, 1997; Smith and Zhang, 1998; U.S. EPA, 1998a). Evidence from human and animal 

studies (McKinney et al., 1991; Shaw et al., 1984; Shu et al., 1988; Buckley et al., 1989) 

suggests that increases in childhood leukemia may be associated with in utero exposures and 

maternal and paternal exposure prior to conception. 

Evidence in animals suggests that exposure to benzene in utero alters maturation of 

lymphocytes, erythrocytes, and granulocytes (OEHHA, 1997). The consequences of in utero 

exposure to benzene can be detected as alterations in cell population numbers and functional 

properties into adulthood. Damage during the initial in utero stages of hematopoiesis could have 

lasting effects, as has been demonstrated for a number of other toxicants (OEHHA, 1997). 

Exposure of mice to benzene in utero during development has been shown to cause 

changes in hematogenic progenitor cells in fetuses, 2-day-old neonates, and 6-week-old adults 

(Keller and Snyder, 1986, 1988; Corti and Snyder, 1996). These results indicate that lasting 

damage to the hematopoietic system can occur during development. It is not known whether the 

effect is the same as in adults or is unique to developmental exposures. Dose-related 

hematotoxicity was observed in the absence of any apparent maternal effects in Swiss-Webster 

mice (Keller and Snyder, 1986, 1988). At high doses, significantly increased skeletal variations 

in CF-1 mouse fetuses (Murray et al., 1979) and in rabbits (Ungvary and Tatrai, 1985) were 

observed. Corti and Snyder (1996) found effects on precursor cells but provided no information 

on peripheral blood effects. The absence of a clear NOAEL for fetal/neonatal exposure makes it 

difficult to directly compare fetal/neonatal and adult toxicity. 

There are few data on the effects of direct exposure of children to benzene. However, 

some indirect evidence suggests that children may be susceptible to benzene-induced 

hematotoxicity. There is mounting evidence that key changes related to the development of 

childhood leukemia occur in the developing fetus. Several studies have reported that genetic 

changes related to eventual leukemia development occur before birth. For example, there is one 

study of genetic changes in sets of twins who developed T cell leukemia at 9 years of age 

(reviewed in Smith and Zhang, 1998). Because of their small size, increased activity, and 

increased ventilation rates, as compared with those of adults, children may have greater exposure 

to benzene in the air on a unit-body-weight basis (U.S. EPA, 1998a). Infants and children also 

109
may be vulnerable to leukemia induction from benzene because their hematopoietic cell 

populations are undergoing maturation and differentiation (U.S. EPA, 1998a). 

4.4.2. Gender Differences 

Human exposure data regarding differential effects of benzene exposure on males and 

females are very limited. Some studies indicate that females may be at greater risk than males. 

Sato et al. (1975) exposed male and female workers to 25 ppm (80 mg/m3) benzene for 2 hours 

and showed that benzene was retained longer in the female subjects. During 5 hours of 

exposure, the blood concentrations were higher in men, whereas at the end, exhalation breath 

levels of both men and women were essentially equivalent. However, 4 hours after the exposure 

had stopped, the female subjects’ blood and exhaled air benzene levels were higher than those of 

male subjects. The shape of the decay curve was significantly steeper in the males. Recently, 

Brown et al. (1998), using the Sato et al. (1975) data, developed a PBPK model simulation for 

adult men and women. Results demonstrated that physiochemical gender differences result in 

women metabolizing 23–26% more benzene than do men when subjected to the same exposure 

conditions. Benzene blood levels were generally higher in men. These results suggest that 

women may be at significantly higher risk for certain effects of benzene exposure. In the 

occupational benzene exposure study by Rothman et al. (1996a), 21 of the 44 workers in both the 

exposed and control groups were female. The study, however, did not indicate that either gender 

was more affected. 

There are limited indications of gender differences in studies with test animals. The most 

frequently observed gender difference is a greater sensitivity of male mice to benzene. Ward et 

al. (1985) reported that microscopic examination revealed that changes in the thymus, bone 

marrow, lymph nodes, spleen, and reproductive organs occurred more often and with greater 

severity in males than in females following inhalation exposure at 300 ppm (958 mg/m3

) of CD-1 

mice for 13 weeks. Male DBA/2 mice showed greater depression in bone marrow cellularity 

than did female DBA/2 mice in response to inhalation exposure to 300 ppm (958 mg/m3

) for 13 

weeks (Luke et al., 1988a). 

Male mice were also more sensitive to benzene by oral exposure (NTP, 1986). In a 17-

week oral gavage study, the LOAEL for hematologic effects in male B6C3F1 mice was 50 

mg/kg, compared with 400 mg/kg for females. Female rats, in contrast, appear to be more 

sensitive than males to benzene toxicity. In the NTP (1986) 17-week oral gavage study, the 

LOAEL for hematologic effects for female F344/N rats was 25 mg/kg, compared with 100 

mg/kg for males. 

Male mice are twofold to threefold more sensitive to the genotoxic effects of benzene, as 

measured by micronuclei induction and sister chromatid exchanges (Luke et al., 1988a). 

Available studies indicate that the differential susceptibility to benzene is subject to hormonal 

regulation of the CYP2E1. Renal levels of CYP2E1 in males are 20-fold higher than in females, 

but studies have indicated that renal levels of CYP2E1 can be induced in females by testosterone 

treatment (Hu et al., 1993; Pan et al., 1992). Kenyon et al. (1996) found that male B6C3F1 mice 

have an almost twofold faster rate of benzene oxidation than do females. Phenol disappearance 

from the blood was also faster in male mice, suggesting that phenol metabolism is faster in 

males. These differences in benzene metabolism correlated with the sensitivity to genotoxicity. 

110
Corti and Snyder (1996) found that CFU-E levels were depressed in livers of male 

fetuses of Swiss-Webster mice exposed in utero to 10 ppm (32 mg/m3

) benzene in comparison 

with controls. Similarly, bone marrow CFU-E was also depressed in 6-week-old adult male 

mice that had been exposed in utero. No depression in CFU-E was observed in female fetuses or 

adult females. This suggests that differences in the sensitivity of male mice can be expressed in 

utero and persist into adulthood. This result is in contrast to the results observed by Siou et al. 

(1981), who found that immature male and female mice did not differ in sensitivity to benzene. 

Green et al. (1978) reported that exposure of pregnant mice to 2200 ppm (7028 mg/m3

) 

benzene caused a greater increase in missing sternebrae in female fetuses than in male fetuses, 

but this was not observed at concentrations of 300 ppm (958 mg/m3) or less. The dose causing 

this effect induced extreme maternal toxicity, complicating interpretation of this effect. The 

apparently increased sensitivity of the female fetus to the effects of benzene is supported by 

other observations that female rabbits, mice, and rats are more sensitive to these effects of 

benzene (Desoille et al., 1961; Ito, 1962a-d; Sato et al., 1975). The investigators tentatively 

suggested that the differences in the responses of the males and females may be related to 

hormonal differences. 

4.4.3. Genetically Susceptible Populations 

Significant sources of variability in the population stem from genetic polymorphisms in 

key enzymes involved in the metabolism of benzene, namely, CYP2E1, NADPH-dependent 

quinone oxidoreductase, MPO, GSH transferase, and others. Dietary and endogenous sources of 

phenol, hydroquinone, and other primary metabolites of benzene confer potentially large 

differences in susceptibility to benzene toxicity. These polymorphisms may increase or decrease 

an individual’s susceptibility to the toxic effects of benzene, as described below. 

Smart and Zannoni (1984, 1985) demonstrated the importance of reductase activity in 

bone marrow as a protective mechanism against benzene toxicity. Recent studies have also 

shown the importance of polymorphic forms of GSH transferase as a protective characteristic. 

By the same token it appears that the reason that the liver is not a target for benzene toxicity is 

its high level of reductase activity, compared with the relatively low level in the bone marrow, 

where oxidative reactions predominate. Furthermore, London et al. (1997) described a 

polymorphism in MPO, an enzyme important in activation of polyhydroxylated metabolites of 

benzene to quinones, suggesting that people who genetically display lower levels of this enzyme 

may display a lower risk to carcinogenesis. Although biomarkers for susceptibility to benzene 

have not been validated, it might be hypothesized that an individual who displays high CYP2E1 

activity and low GSH transferases in liver and/or bone marrow and low bone marrow reductase 

and high MPO could be highly susceptible to benzene toxicity. 

Ross et al. (1996) have suggested that NAD(P)H:quinone oxidoreductase (NQO1 

NAD(P)H) may play a critical protective role in benzene toxicity. This enzyme detoxifies the 

reactive 1,4-benzoquinone generated by MPO by reducing it back to hydroquinone. Ross (1996) 

summarized evidence suggesting that the target cells of benzene toxicity have a high ratio of 

MPO:NQO1 activities in vivo. Traver et al. (1992) characterized a point mutation in the NQO1 

gene that leads to a total loss of NQO1 activity. This appears to be a true polymorphism in the 

NQO1 gene, because Rosvold et al. (1995) found that the frequency of this mutant allele in a 

111
reference population was 13% and Edwards et al. (1980) reported that NQO1 was absent in 4% 

of samples taken from a British population. If the postulated role of NQO1 in detoxification of 

benzene metabolites is correct, then individuals who lack this enzyme activity could be 

especially susceptible to benzene toxicity. Rothman et al. (1997) carried out a case-control study 

in a population of benzene-exposed workers in Shanghai, China. CYP2E1 and NQO1 genotypes 

were determined by polymerase chain reaction-restriction fragment length polymorphism 

analysis, and CYP2E1 activity was estimated by the fractional excretion of chlorzoxanone. 

Investigators found that subjects who had both rapid chlorzoxanone excretion and two copies of 

the NQO1 609CT mutation had a 7.6-fold increased risk of hematotoxicity when compared with 

subjects who had slow chlorzoxanone excretion and who carried one or two wild-type NQO1 

alleles. 

The joint effects of CYP2E1 and NQO1 status, shown in Table 9, indicate that rapid 

CYP2E1 activity or a nonfunctional NQO1 increase an individual’s risk of benzene 

hematotoxicity. Those individuals who had rapid CYP2E1 activity and heterozygous or 

homozygous variant NQO1 genotypes were at greatest risk of developing benzene-induced 

hematotoxicty. 

Seaton et al. (1994) reported differences of up to 13-fold in liver microsome CYP2E1 

among individual humans. Hepatic metabolism of benzene by the CYP2E1 enzyme has been 

shown to be the first step in generation of reactive benzene metabolites that are responsible for 

the toxicity of benzene. Thus, differences in CYP2E1 between individual humans could indicate 

potential differential susceptibility to benzene toxicity. 

Rossi et al. (1999) investigated the role of genetic polymorphisms in modulating urinary 

excretion of two benzene metabolites—MA and S-phenylmercapturic acid—in 59 nonsmoking 

city bus drivers, who were professionally exposed to benzene via vehicle exhausts. Genetic 

Table 9. Joint effects of CYP2E1 activity and NQO1 genotype on benzene-induced 

hematotoxicity in Chinese Workers 

CYP2E1 activity NQO1 genotype 

Slow Wild type 

Slow Variant 

Rapid Wild type 

Rapid Variant 

Cases 

8 

6 

21 

13 

Odds ratio (95% CI) 

benzene hematotoxicitya 

1.0 

2.4 (0.6–9.7) 

2.9 (1.0–8.2) 

7.6 (1.8–31.2) 

a

Adjusted for the matching variables age and sex. 

Source: Rothman et al., 1997. 

polymorphisms at six loci encoding cytochrome P450-dependent monooxygenases (CYP2E1 and 

CYP2D6), GSH transferases (GSTT1, GSTP1, and GSTM1), and NQOR were determined by 

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polymerase chain reaction based methods. Metabolic variability was observed for CYP2D6, 

GSTT1, and NQOR, potentially contributing polymorphism as one of the risk factors for 

benzene-induced adverse health effects. However, no evidence emerged for a possible role of 

CYP2E1, GSTM1, and GSTP1 polymorphisms in determining the wide differences observed in 

the biotransformation. 

4.5. HAZARD IDENTIFICATION SUMMARY 

Studies in experimental animals and human occupational studies have pointed to the 

changes in peripheral blood and bone-marrow-induced exposure to benzene as being among the 

most sensitive indicators of the compound’s toxicity. Principal among the adverse effects 

described are leukopenia, lymphocytopenia, granulocytosis, anemia, and reticulocytosis. 

Human occupational exposure studies have revealed clear evidence of benzene-induced 

hematotoxicity (Table 1). Difficulties arise in directly applying most of these findings to 

quantitative estimation of potential human health impacts. This is due to uncertainty in defining 

the level of benzene exposure and to potential exposure to mixtures of other potentially harmful 

substances. One human occupational study, however, reported evidence that can be used to 

support the derivation of RfC and RfD values for benzene. In a small cross-sectional study of 44 

age- and gender-matched controls, Rothman et al. (1996a) observed a dose-response relationship 

and inverse correlation between hematologic responses and exposure level. Six blood 

parameters (ALC, WBC count, RBC count, HCT, platelets, and MCV) were significantly 

different in the high-benzene exposure group (8-hour TWA of 91.9 ppm [294 mg/m3

]) in 

comparison with controls. However, in the lower-exposure group (median 8-hour TWA of 13.6 

ppm [43.4 mg/m3

], only ALC, RBCs, and platelets were significantly different. 

In a second case-control study of a rubber worker cohort employed between 1939 and 

1976 that relied on controversial exposure estimates (Ward et al., 1996), a strong exposureresponse relationship between WBC count and estimated benzene exposure concentration for 30, 

90 and 180 days before the blood test date was evident; however, there was only a weak positive 

exposure response for RBC count. The maximum daily exposure estimate was 34 ppm (109 

mg/m3

). There was no evidence for a threshold, suggesting that exposure to relatively low levels 

of benzene (e.g., < 5 ppm [16 mg/m3

]) could result in hematologic suppression, if the exposure 

estimates are correct. 

Further support for the hypothesis that low exposures to benzene can reduce lymphocyte 

counts comes from a recent study (Bogadi-Sare et al., 2000) of female employees in the 

shoemaking industry. This study found that the number of circulating B lymphocytes was lower 

in 49 shoe workers exposed to benzene concentrations lower than 15 ppm (48 mg/m3

) as 

compared with 27 nonexposed controls. The authors concluded that benzene concentrations 

lower than 15 ppm (48 mg/m3) can induce depression of circulating B lymphocytes. Additional 

evidence that low exposures to benzene induce reductions of blood parameters comes from a 

study of workers employed in factories in China (Dosemeci et al., 1996). Significant relative 

risks of “benzene poisoning” were recorded in workers exposed to as little as 5–19 ppm (16–61 

mg/m3

) during the most recent 1.5 years as compared with workers exposed to < 5 ppm (16 

mg/m3

). The authors defined benzene poisoning by two criteria: a WBC count of < 4000 

cells/mm3

 and a platelet count of < 80,000/mm3. These data suggest that a threshold of 

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hematotoxicity in humans is observable in the 5–19 ppm (16–48 mg/m3

) range of exposure to 

benzene. 

Hematotoxicity has been observed in experimental animal studies following both oral and 

inhalation exposure (Table 6). Hematologic abnormalities developed rapidly and were evident 

after only 2 weeks of exposure, but there was comparatively little progression with longer 

exposure periods (Ward et al., 1985; Cronkite et al., 1985). The dose-response studies of Ward 

et al. (1985) identified a NOAEL of 30 ppm (96 mg/m3

) and a LOAEL of 300 ppm (958 mg/m3) 

in CD-1 mice. The results of Cronkite et al. (1985, 1989) suggest that peripheral blood elements 

may have the capacity to recover from benzene-induced toxicity, depending on the duration of 

exposure and concentration. For example, lymphocyte counts, which were depressed in 

C57BL/6 mice exposed to benzene at 300 ppm (958 mg/m3

) for 2–16 weeks, tended towards 

normal values during a 4- to 8-week recovery period. 

Benzene-induced peripheral blood abnormalities reflect a disruption at all levels of 

hematopoiesis in the bone marrow (Toft et al., 1982; Snyder, 1987; Cronkite et al., 1985; Seidel 

et al., 1989; MacEachern et al., 1992). Bone marrow cellularity, stem cell compartments, 

granulocytic and erythropoietic progenitor cells, and bone marrow macrophages have all been 

observed to be adversely affected following benzene exposure. Recent evidence has suggested 

that the inhibitory effect of benzene on lymphocyte proliferation can be mediated through the 

suppression of cytokine production (Renz and Kalf, 1991). The fact that both in vivo and in 

vitro studies have demonstrated the ability of some metabolites to induce toxicological effects 

qualitatively similar to those of benzene provides evidence consistent with the concept that the 

hematotoxicity of benzene may be mediated through its metabolites. 

There is no convincing evidence that benzene produces either reproductive or 

developmental toxicity in humans. The available data are summarized in Tables 4 and 5. Most 

studies consisted of small numbers of subjects, lacked important experimental details, involved 

(in almost all cases) concomitant exposure to other chemicals, and did not provide monitoring 

data or quantitative dose-response information. 

A number of reproductive and developmental studies have been conducted in a wide 

range of experimental animals (Tables 7 and 8). Changes in testicular weight in guinea pigs 

(with slight histopathologic alterations) and rats were observed by Wolf et al. (1956), and 

reductions in ovarian and testicular weights with histopathologic alterations were observed in 

mice by Ward et al. (1985) after prolonged exposure to benzene. However, the concentrations 

were higher than those observed to cause hematotoxicity. Several studies have suggested that 

benzene can cause developmental toxicity in the absence of maternal toxicity in rats. Effects 

observed include reduced body and liver weights in weanlings (Kuna et al., 1992), skeletal 

variations in fetuses (Kuna and Kapp, 1981), and reduced fetal weight and delayed ossification 

in fetuses (Green et al., 1978; Coate et al., 1984; Ungvary and Tatrai, 1985). Although benzene 

can cause reproductive and developmental toxicity, the LOAEL and NOAEL values associated 

with reproductive/developmental toxicity are higher than for hematotoxicity. 

The hematologic effects in offspring of dams exposed to benzene during gestation were 

observed at exposure concentrations similar to those in adult animals (Keller and Snyder, 1986, 

1988; Corti and Snyder, 1996). Changes in granulopoiesis and hematopoiesis in offspring at 

114
some growth stages were observed. Thus, the results were not consistent. The results of these 

studies cannot be used for quantitative evaluation of human health risk because of the limited 

number of replicate animals. 

Symptoms of neurotoxicity have been observed in humans following acute exposure to 

relatively high concentrations of benzene (Brief et al., 1980; Kraut et al., 1988). Kahn and 

Muzyka (1973) reported subjective symptoms—headache, fatigue, difficulty sleeping, and 

memory loss—in factory workers exposed to benzene. Baslo and Aksoy (1982) reported 

neurological abnormalities in patients diagnosed with aplastic anemia after prolonged exposure 

to benzene. These reports, however, have obvious deficiencies: lack of exposure data, small 

numbers of subjects, and unknown exposure to other chemicals. Thus, no reliable quantitative 

evaluation of neurotoxicity in humans is possible. 

Neurotoxicity studies with experimental animals were also very limited. The exposure 

levels were generally high and exposure durations short. The longest exposure period was 4 

weeks (Hsieh et al., 1988a). These investigators observed significant increases in monoamine 

neurotransmitters in the brain of rats at doses of 8 and 40 mg/kg/day, but further increases were 

not observed at 180 mg/kg/day. Benzene-induced behavioral and learning disorders in mice 

following inhalation exposures of 100 ppm (319 mg/m3

) or greater for 5 days (Evans et al., 1981; 

Dempster et al., 1984). Li et al. (1992) observed increased grip strength and enhanced rapidresponse maze performance after exposure to 0.78 ppm (2.5 mg/m3

) for 30 days and decreases in 

these parameters at higher concentrations. There were, however, several limitations in the 

experimental procedures used in these experiments. There is a limited body of evidence 

indicating that benzene is neurotoxic; however, there are no human or animal studies that could 

be used for quantitative evaluation of potential human health risks. 

No comprehensive immunotoxicity studies in human populations have been reported, but 

the study by Rothman et al. (1996a) indicates decreased ALC as the most sensitive indicator of 

benzene toxicity. This could be interpreted as an effect on immune function. 

Immunotoxicity studies in experimental animals have demonstrated dose-related adverse 

effects on spleen weight and cellularity and on several measures of immune function following 

both oral and inhalation exposure to benzene (Hsieh et al., 1988b; Fan, 1992; White et al., 1984; 

Rosenthal and Snyder, 1985; Rozen et al., 1984). NOAEL and LOAEL values for adverse 

effects on immune function were similar to those established for hematotoxic effects. These 

studies indicate that even short-term exposures to benzene adversely affect the immune response 

in experimental animals. 

There is no convincing evidence to indicate that children are more susceptible to the toxic 

effects of benzene; however, there is evidence that differences in gender and subpopulation 

susceptibility may exist. Differences in responsiveness to benzene have been observed among 

species. Mice were found to be more sensitive than rats (Ward et al., 1985; Snyder et al., 1978, 

1980). Intraspecies variability has also been demonstrated. AKR/J mice were found to be more 

sensitive than C57BL/6 mice (Snyder et al., 1978, 1980). Absorption studies with humans 

suggest that absorption by females is higher than by males, and modeling results indicate that 

females metabolize 23–26% more benzene than men under the same exposure conditions (Sato 

et al., 1975; Brown et al., 1998). Differences in benzene metabolism could result in differences 

115
in susceptibility, because conversion of benzene to metabolites is necessary for expression of 

toxicity. Several studies have indicated that male mice are more sensitive than females (Ward et 

al., 1985; NTP, 1986; Siou et al., 1981; Kenyon et al., 1996; Corti and Snyder, 1996). In 

contrast, other studies using mice, rats, and rabbits have indicated that females are more sensitive 

than males (Desoille et al., 1961; Ito, 1962a-d; Sato et al., 1975). 

The balance of NQO1 NAD(P)H activity to peroxidase activity in cells may be important 

in determining benzene toxicity by modulating the concentration of 1,4-benzoquinone, a 

suspected toxic metabolite. A mutant allele for lack of NQO1 NAD(P)H activity has been 

reported in a human population at a frequency of 13% (Rosvold et al., 1995). Edwards et al. 

(1980) reported that 4% of the British population lack NQO1 NAD(P)H activity. Rothman et al. 

(1997) also observed that Chinese workers homozygous for the mutant allele and with high 

chlorzoxazone excretion (a measure of CYP2E1 activity) had a 7.6-fold higher risk of 

developing benzene toxicity. Thus, there is good experimental evidence to indicate that 

benzene-sensitive human subpopulations may exist. 

Several experimental animal studies have observed that ethanol and benzene have an 

interactive effect on the production of hydroxylated benzene metabolites (Nakajima et al., 1987). 

When administered at the same time, ethanol and benzene enhanced the number of erythroid 

progenitor cells in C57BL/6J mice (Baarson and Snyder, 1991), decreased the number of CFU-E 

cells per femur in BDF1 mice (Seidel et al., 1990), and produced a 70% reduction of CFU-E in 

male Swiss-Webster mice livers (Corti and Snyder, 1996). On the other hand, a 40% elevation 

of CFU-E occurred in female Swiss-Webster mice exposed to benzene and ethanol compared to 

unexposed controls (Corti and Snyder, 1996). There are no relevant published data on the 

interactive effects of benzene exposure and ethanol use in the human population. 

5. DOSE-RESPONSE ASSESSMENTS 

The primary target organs for benzene toxicity are the hematopoietic, immune, and 

nervous systems. There are several reports of short-term exposure in humans and animals. Data 

on the effects of benzene on the human hematopoietic system following occupational inhalation 

exposure are scant, but they indicate effects such as leukopenia, anemia, and thrombocytopenia. 

Data on hematologic effects in experimental animals from short-term inhalation/oral exposures 

are extensive, indicating changes in peripheral erythrocytes (e.g., Cronkite et al., 1985; Rozen et 

al., 1984), in peripheral leukocytes (e.g., Aoyama, 1986; Li et al., 1986; Green et al., 1981a), and 

in bone marrow cells (e.g., Cronkite et al., 1989; Dempster and Snyder, 1991). However, these 

studies were generally determined to be unsuitable for derivation of the RfD/RfC because they 

involved short-term exposures and small numbers of experimental animals; in addition, exposure 

levels were often too high to establish meaningful LOAELs or NOAELs. 

Data on adverse hematologic effects in humans following intermediate and chronic 

duration exposures to benzene in occupationally exposed individuals are available (Aksoy et al., 

1971, 1972, 1974, 1987; Cody et al., 1993; Greenberg et al., 1939; Kipen et al., 1989; Li et al., 

1994; Townsend et al., 1978; Yin et al., 1987b). However, in the majority of these studies the 

exposure level estimates and duration of exposure are poorly defined and/or characterized, and 

potential confounding is problematic; therefore, these studies could not be used in determinating 

116
the RfC and RfD, with the exception of the Rothman et al. (1996a) study, which is used as the 

key study. The study of occupational inhalation exposure to benzene by Tsai et al. (1983) seems 

to suggest a NOAEL for hematologic effects; however, the study had no appropriate screening 

guidelines or referral criteria with which to contrast results, no follow-up, a very pronounced 

healthy-worker effect, and a very young employed survivor population. No relevant human data 

are available to evaluate hematologic effects following oral exposure. 

Several subchronic and chronic studies in experimental animals following inhalation 

exposure (see Table 6) and oral exposure (NTP, 1986; Huff et al., 1989; Wolf et al., 1956; see 

also Section 4.4.2.1) identified NOAELs and LOAELs for hematologic effects and were 

considered in the derivation of the RfC and RfD. Ultimately, the results of the Ward et al. 

(1985) subchronic inhalation study and the NTP (1986) chronic gavage study were selected as 

the most appropriate animal data for deriving an RfC and an RfD. These RfC and RfD values 

are calculated for comparison with the values derived from the Rothman et al. (1996a) human 

study. 

Whether the hematotoxic/immunotoxic effects of benzene and its carcinogenic effects are 

due to a common mechanism is not yet known. This is in part due to the fact that although the 

bone marrow depressive effects of benzene in humans can be readily duplicated in several 

experimental animal model systems, a suitable experimental animal system for the induction of 

leukemia has not yet been developed. In addition, the hematotoxicity/immunotoxicity of 

benzene leads to significant health effects apart from potential induction of leukemia. Although 

the decreased ALC and leukemias observed from benzene exposure both result from bone 

marrow toxicity, they do not necessarily result from the same mechanisms, and decreased ALC 

may not be a necessary precursor for leukemia. For example, decreased ALC may be the result 

of cytotoxicity independent of the interaction of benzene metabolites with DNA and/or DNAassociated proteins and tumor formation. Thus, integration of the cancer and noncancer health 

assessments of benzene is not considered possible given the current state of knowledge regarding 

mode of action. 

5.1. INHALATION REFERENCE CONCENTRATION (RfC) 

5.1.1. Choice of Principal Study and Critical Effect 

The human occupational inhalation study by Rothman et al. (1996a) was selected as the 

principal study for the derivation of the RfC because it is a well-conducted human exposure 

study that demonstrates a dose-response relationship for hematologic effects, responses that are 

considered to be among the more sensitive indices of benzene toxicity. The California EPA 

(2001) selected the Tsai et al. (1983) study to develop a drinking water noncancer protective 

concentration of 0.026 mg/L, because no increased hematological changes were observed among 

U.S. refinery workers chronically exposed to an average 0.53 ppm of benzene. This value was 

assumed to be a NOAEL by the California EPA. The U.S. EPA did not choose this study to 

develop an RfC or an RfD because of deficiencies in the study design. However, the California 

EPA’s numbers were similar to those of the U.S. EPA, although different studies were used. The 

cross-sectional Rothman et al. (1996a) study provides exposure-response data for some of the 

lowest exposure concentrations at which effects have been observed, as discussed in Section 

4.1.2.1. These exposure-response data are suitable for BMD modeling to derive a point of 

117
departure for the calculation of the RfC. Reduced ALC is the most sensitive of the hematologic 

effects reported in this study, and it was selected as the critical effect. For comparison purposes, 

a chronic inhalation RfC is also derived from the subchronic experimental animal study of Ward 

et al. (1985), based on the critical effect of decreased HCT (see Section 5.1.5). 

Rothman et al. (1996a) compared the hematologic outcomes of 44 workers 

occupationally exposed to benzene with those of 44 age- and gender-matched unexposed 

controls, all based in Shanghai, China. ALC, WBCs, RBCs, and platelets were all significantly 

decreased and MCV was significantly increased in the exposed group (median 8-hour TWA of 

31 ppm [99 mg/m3

]) as compared to those in the control group. These effects are consistent with 

the hematotoxic effects of benzene shown in Aksoy (1989), Goldstein (1988), and Dosemeci et 

al. (1997). In the low-dose subjects (< 31 ppm [99 mg/m3

], median 8-hour TWA of 13.6 ppm 

[43.4 mg/m3]), the ALC, RBCs, and platelet count were reduced compared with controls. 

Similarly, in a selected subgroup exposed to a median 8-hour TWA of 7.6 ppm (24 mg/m3

) 

benzene, a statistically significant difference in ALC versus controls was observed (p<0.03). 

The study by Rothman et al. (1996a) is notable among epidemiology studies because 

benzene exposures were monitored; exposure to other chemicals, including toluene, was 

minimal; and subjects were compared with matched controls. Furthermore, a dose-response 

relationship was established between ALC and benzene exposure, as monitored by organic vapor 

passive dosimetry and the level of benzene metabolites in the urine. As discussed in Section 

4.1.2.1, the median 8-hour TWA of 7.6 ppm (24 mg/m3

) was designated the LOAEL for these 

effects and is used herein to calculate a chronic inhalation RfC for benzene hematotoxicity in 

humans for comparison with the RfC derived by BMD modeling of the ALC exposure-response 

data. 

The choice of a reduction in ALC as the primary effect in an RfD/RfC derivation is 

partially based on the response’s potential role as a “sentinel” effect for a cascade of early 

hematologic and related biological changes that might be expected to result in the more profound 

examples of benzene poisoning observed in other cohorts of the National Cancer 

Institute/Chinese Academy of Preventive Medicine study, as described by Dosemeci et al. 

(1996). Considered together, the statistically significant inverse correlation between ALC and 

the level of benzene metabolites in the urine of exposed subjects in the Rothman et al. (1996a) 

study and the strong relationship between exposure to benzene in persons with benzene 

poisoning observed by Dosemeci et al. (1996) point clearly to the utility of fluctuations in ALC 

as a marker for the onset of potentially harmful hematologic changes. That ALC depletion is 

accompanied by gene-duplicating mutations in somatic cells under the same range of exposure 

conditions suggests that benzene can cause repeated damage to longer-lived stem cells in human 

bone marrow, further implicating the compound as etiologically important in the onset of 

benzene-associated leukemia. These findings underline the importance of basing public health 

concern for benzene on a toxicological effect that is representative of the earliest biological 

changes induced by the compound. 

In summary, there is overwhelming evidence, in both experimental animals and humans, 

in the published literature cited in this report and in the cancer update documents (U.S. EPA, 

1998a, 1999a), that chronic exposure to benzene leads to the onset of irreversible bone marrow 

118
depression, which is characterized by anemia, leukocytopenia with emphasis on 

lymphocytopenia, and/or thrombocytopenia. 

It is a well-known fact that benzene is a leukemogen. The mechanistic steps that lead to 

that outcome involve decreases in circulationg blood cells, as characterized by abnormal marrow 

architecture, inadequate hematopoiesis, and demonstration of chromosomal damage in many 

cells. Benzene toxicity is a continuum of events, leading from decreases in circulating blood 

cells to pancytopenia and aplastic anemia or to MDS and acute nonlymphocytic leukemia. 

Although the role of several metabolites of benzene and the “dose x times” relationship are only 

partially understood, the ultimate result as a function of marrow dysplasia or aplasia or leukemia 

is undisputable. In the Rothman et al. (1996a) study, benzene air levels were inversely 

correlated with the absolute lymphocyte counts among exposed workers. This has been the 

subject of more detailed discussion in several sections of this document as well as in the cancer 

update documents (U.S. EPA, 1998a, 1999a). The reader’s attention is drawn to sections 4.1, 

4.2, 4.5, as well as to some selected references: Rothman et al. (1996a, b), Dosemeci et al. 

(1997), Bogardi-Sari et al. (2000), Hsieh et al. (1988b), NTP (1986), and Ward et al. (1985). 

5.1.2. Benchmark Dose Modeling 

The exposure-response data from Rothman et al. (1996a, Table IV) that were used for the 

BMD modeling are reproduced below. 

Median exposure ALC Transformed 

(ppm; 8-hr TWA) Number of subjects (mean ± SD × 103

/μL blood) exposures 

0.02 (control) 44 1.9 ± 0.4 0.0198 

13.6 22 1.6 ± 0.3 2.68 

91.9 22 1.3 ± 0.3 4.53 

The modeling was done using EPA’s Benchmark Dose Modeling Software (version 

1.20). The data are fairly supralinear, that is, the change in ALC per unit change in exposure 

decreases with increasing exposure; therefore, in order to fit the data with one of the available 

continuous models, the exposure levels were first transformed according to the equation 

d1 = ln(d + 1). Then the exposure-response data were fit using the continuous linear model, 

which has the form 

Y[dose] = 0 + 1*dose, 

where, in this case, Y[dose] is the mean ALC and dose is the transformed exposure, d1. The 

parameters were estimated using the method of maximum likelihood. A constant variance model 

was used. The resulting parameter estimates are: 

Variable Estimate SE 

119
0 1.91029 26.8966 

1 -0.129822 70.7846 

The model fit was good, as is apparent from the p-value of 0.5443 and the graphical display in 

Figure 2. 

In the absence of a clear definition for an adverse effect for this continuous endpoint, a 

default benchmark response (BMR) of one standard deviation change from the control mean was 

selected, as suggested in EPA’s draft Benchmark Dose Technical Guidance Document (U.S. 

EPA, 2000b). This default definition of a benchmark response for continuous endpoints 

corresponds to an excess risk of approximately 10% for the proportion of individuals below the 

2nd percentile (or above the 98th percentile) of the control distribution for normally distributed 

effects (see U.S. EPA, 2000b). The software uses the estimated standard deviation. For the 

resulting BMC, a 95% lower confidence limit (BMCL) was calculated using the likelihood 

profile method. A BMC of 2.69 and a BMCL of 2.10 were obtained with the transformed 

exposures. Transforming these values back to the original exposure scale yields the following 

values: BMC = 13.7 ppm, 8-hour TWA; BMCL = 7.2 ppm, 8-hour TWA. 

A two-degree restricted polynomial model also adequately fits the data, yielding a BMC 

of 3.00 and a BMCL of 2.14 with the transformed doses (i.e., 19.1 and 7.5 ppm 8-hour TWA, 

respectively); however, the linear model was selected because it is the most parsimonious (i.e., it 

120
Figure 2. Figure 2. Linear model of ALC data from Rothman et al. (1996a). Linear model of ALC data from Rothman et al. (1996a). 

is the model with the fewest parameters that fits the data) and 2 × (the absolute difference in the 

log-likelihood values between the two models) < 31

2

, that is, the parameter 2 for the coefficient 

of dose2

 in the two-degree polynomial model is not statistically different from 0. Results using a 

restricted power model were identical to those using the two-degree polynomial. 

Note that the 7.6 ppm exposure subgroup (11 subjects) was not included explicitly in the 

exposure-response modeling because it was a subset of the 13.6 ppm group (22 subjects) and the 

data were not available for the remaining 11 subjects in the 13.6 ppm group (nor were the 

individual data available). Thus, using the 7.6 ppm group rather than the 13.6 ppm group would 

have meant that some of the data, for one-fourth of the exposed subjects, would not have been 

used at all. Comparison analyses were conducted to examine the effects of using the 7.6 ppm 

group and either omitting the subjects in the 2nd exposure quartile or estimating the exposure 

level of the 2nd quartile as the midpoint between 13.6 and 31 ppm. The BMCLs were similar to 

those obtained when the 7.6 ppm group was not included separately (see Table 10). 

5.1.3. RfC Derivation 

As suggested in the draft BMD technical guidance document (U.S. EPA, 2000b), the 

BMCL is chosen as the point of departure for the RfC derivation. An adjusted BMCL is 

calculated by converting ppm to mg/m3

 and adjusting the 8-hour TWA occupational exposure to 

Table 10. Results of BMC modeling of Rothman et al. (1996a) data on benzene and 

ALCa 

121
Dataset BMC (ppm) BMCL (ppm) 

without 7.6 ppm subgroup 13.7 7.2 

with 7.6 ppm subgroup, 

omitting 2nd quartile 

14.7 7.4 

with 7.6 ppm subgroup, 

estimating 2nd quartile 

17.1 8.4 

a

Using log-transformed exposures, a linear model, and a BMR level of one standard deviation change from the 

control mean. 

an equivalent continuous environmental exposure. The BMCL is first converted to mg/m3 using 

the molecular weight of 78.11 for benzene and assuming 25°C and 760 mm Hg: 

BMCL (mg/m3) = 7.2 ppm × 78.11/24.45 = 23.0 mg/m3 

The converted value is then adjusted from the 8-hour TWA to an exposure concentration 

adjusted for continuous exposure using the default occupational minute volume (U.S. EPA, 

1994). 

BMCLADJ = BMCL (mg/m3

) × (VEho/VEh) × 5 days/7 days 

BMCLADJ = 23.0 mg/m3

 × (10 m3/20 m3) × 5 days/7 days = 8.2 mg/m3 

where: 

BMCLADJ = the BMCL dosimetrically adjusted to account for continuous exposure 

BMCL = occupational exposure level (8-hour TWA) 

VEho = human occupational default minute volume (10 m3/8 hours) 

VEh = human ambient default minute volume (20 m3/24 hours). 

To calculate an RfC using the BMCLADJ value of 8.2 mg/m3, several UFs were applied. 

First, because the BMC is considered to be an adverse-effect level, an effect-level 

extrapolationfactor analogous to the LOAEL-to-NOAEL UF is used. EPA is planning to 

develop guidance for applying an effect-level extrapolation factor to a BMD. In the interim, a 

factor of 3 is used in this analysis, based on the professional judgement that, while the BMC 

corresponds to an adverse-effect level at the low end of the observable range, the endpoint is not 

very serious in and of itself. Decreased ALC is a very sensitive sentinel effect that can be 

measured in the blood, but it is not a frank effect, and there is no evidence that it is related to any 

functional impairment at levels of decrement near the BMR. For a more serious effect, a larger 

factor, such as 10, might be selected. 

122
Second, a factor of 10 was used for intraspecies differences in response (human 

variability) as a means of protecting potentially sensitive human subpopulations. Third, a 

subchronic-to-chronic extrapolation factor was applied because the mean exposure duration for 

the subjects in the principal study was 6.3 years, which is less than the exposure duration of 7 

years (one tenth of the assumed human life span of 70 years) which has been used by the 

Superfund program as a cut-off for deriving a subchronic human reference dose (U.S. EPA, 

1989). However, because the mean exposure duration (range 0.7 to 16 years) was near the 

borderline of what would be considered chronic (i.e., 6.3 years vs. 7 years), a partial value of 3 

(vs. 10) was felt to be appropriate for the UF. 

Finally, a UF of 3 was chosen to account for database deficiencies because, despite the 

extensive nature of the overall toxicological database for benzene, no two-generation 

reproductive/developmental toxicity study is available. Therefore, an overall UF of 3 × 10 × 3 × 

3 = 300 is used to calculate the chronic inhalation RfC, as follows: 

RfC = BMCLADJ ÷ UF = 8.2 mg/m3

 ÷ 300 = 3 × 10-2 mg/m3 

5.1.4. Comparison Analysis Based on the LOAEL 

A median 8-hour TWA concentration of 7.6 ppm was designated the LOAEL for the 

Rothman et al. (1996a) study for a comparison calculation of a chronic inhalation RfC for the 

critical effect of reduced ALC. A LOAELADJ can be calculated, as described for the BMCLADJ in 

Section 5.1.3, by converting to mg/m3 and adjusting from the occupational ventilation rate and 

intermittent work-week schedule to a continuous 24-hour exposure, 7 days/week: 

LOAEL (mg/m3) = 7.6 ppm × 78.11/24.45 = 24.3 mg/m3 

LOAELADJ = 24.3 mg/m3 × (10 m3/20 m3) × 5 days/7 days = 8.7 mg/m3 

To calculate an RfC using the LOAELADJ value of 8.7 mg/m3, the following UFs were 

selected. A factor of 10 was applied to account for using a LOAEL because of the lack of an 

appropriate NOAEL. A factor of 10 was used for intraspecies differences in response (human 

variability). A partial UF of 3 for subchronic-to-chronic extrapolation was applied and a UF of 3 

was chosen to account for database deficiencies, as discussed in Section 5.1.3 above. Therefore, 

an overall UF of 10 × 10 × 3 × 3 = 1000 is used in the calculation of the chronic inhalation RfC 

from the LOAEL, as follows: 

RfC = LOAELADJ ÷ UF = 8.7 mg/m3

 ÷ 1000 = 9 × 10-3 mg/m3 

This value is in good agreement with the RfC of 3 × 10-2 mg/m3 calculated from the BMC. 

5.1.5. Comparison Analysis Based on the Ward et al. (1985) Experimental Animal Study 

A chronic inhalation RfC was also derived from the subchronic experimental animal 

study of Ward et al. (1985) for comparison with the RfC of 3 × 10-2 mg/m3

 based on the 

Rothman et al. (1996a) human study. The Ward et al. (1985) study was selected because it used 

a relatively long inhalation exposure duration and an adequate number of animals, and it 

123
provided dose-response data. The study identified both a LOAEL of 300 ppm and a NOAEL of 

30 ppm. The investigators exposed male and female CD-1 mice and Sprague-Dawley rats to 0, 

1, 10, 30, or 300 ppm (0, 3.2, 32, 96, or 960 mg/m3

) benzene 6 hours/day, 5 days/week for 91 

days and measured various hematologic endpoints. The male mouse appears to be the most 

sensitive sex/species in this study. The exposure-response relationships for the different 

hematologic endpoints were modeled using a BMD modeling approach, and decreased HCT was 

chosen as the critical effect. The exposure-response data from Ward et al. (1985, Table II, 91 

days) for the male mouse are reproduced below. 

Exposure Number of % HCT 

(ppm) subjects (mean ± SD) 

0 20 41.1 ± 1.61 

1 20 38.4 ± 3.93 

10 20 40.8 ± 3.22 

30 20 38.4 ± 5.65 

300 20 27.9 ± 4.75 

EPA’s Benchmark Dose Modeling Software (version 1.20) was used for the modeling. 

An assumption of constant variance was used, although the test for homogeneity of the variances 

failed. The continuous linear, polynomial, and power models all resulted in the same BMC and 

BMCL estimates; however, the linear model had better results for the fit statistics. The linear 

model had a p-value of 0.09, which is of borderline adequacy (EPA’s draft Benchmark Dose 

Technical Guidance Document [U.S. EPA, 2000b] recommends a p-value of  0.1), whereas the 

other models had p-values of 0.04. Thus the continuous linear model was selected. This model 

has the form 

Y[dose] = 0 + 1 * dose 

where, in this case, Y[dose] is the mean HCT level and dose is the experiment exposure 

concentration. The parameters were estimated using the method of maximum likelihood. The 

resulting parameter estimates are 

Variable Estimate SE 

0 40.0962 2.44042 

1 -0.0407075 329.232 

See Figure 3 for a graphical display of the data and model. 

In the absence of a clear definition for an adverse effect for this continuous endpoint, a 

default BMR of one standard deviation from the control mean was selected, as suggested in the 

EPA’s draft BMD technical guidance document (U.S. EPA, 2000b). The software uses the 

estimated standard deviation. A 95% lower confidence limit (BMCL) on the resulting BMC was 

calculated using the likelihood profile method. A BMC of 100.7 ppm and a BMCL of 85.0 ppm 

124
Figure 3. Linear model of HCT data from Ward et al. (1985) 

were obtained. 

It should be noted that the dose spacing in this study was less than ideal. Responses in the 

three lower-exposure groups for all the hematologic endpoints tended to clump near control 

group levels, and significant deviations in response were generally seen only in the 300 ppm 

group, with a large exposure range in between, including where the BMC is located, for which 

there are no response data (see Figure 3). Therefore, there is some uncertainty about the actual 

shape of the dose-response curve in the region of the BMR and, thus, some corresponding 

uncertainty about the values of the BMC and BMCL estimates. 

ALCs were not reported in Ward et al. (1985), so this endpoint could not be compared 

with the human ALC results. Total WBC counts were reported, and they exhibited the largest 

percent change in response between the control and the 300 ppm group. However, the data for 

this endpoint also had substantial variance, and because the BMR used for this analysis is a 

function of the standard deviation, WBC counts did not yield the lowest BMC estimate. The 

actual lowest BMC estimates were obtained for increased mean cell Hgb (78 ppm; BMCL = 67 

ppm) and increased mean cell volume (79 ppm; BMCL = 68 ppm); however, these endpoints are 

probably not adverse per se. On the other hand, they are likely to be compensatory effects—and 

thus markers of toxicity—and one could probably justify using them as the critical effects. In any 

event, the BMC estimates are not much different from the BMC of 100 ppm obtained for 

decreased HCT. The results are also similar for total blood Hgb (BMC = 104 ppm; BMCL = 88 

125

ppm). RBC results were in between those for MCV and mean corpuscular Hgb (MCH) and 

those for HCT and total Hgb; however, the model fits were not adequate for the RBC data; thus, 

the RBC results have more uncertainty. 

To derive the RfC, the BMCL is used as the point of departure, as suggested in the draft 

BMD technical guidance document (U.S. EPA, 2000b). An adjusted BMCL is calculated by 

converting ppm to mg/m3

 and adjusting from animal experiment exposure to equivalent 

continuous exposure. (For conversion of the inhalation exposures across species, ppm 

equivalence was assumed; this is identical to using EPA’s inhalation dosimetry methodology 

with the regional gas dose ratio for the respiratory tract region (RGDRr

) = 1 [U.S. EPA, 1994].) 

The BMCL is first converted to mg/m3

 using the molecular weight of 78.11 for benzene and 

assuming 25°C and 760 mm Hg: 

BMCL (mg/m3

) = 85.0 ppm × 78.11/24.45 = 272 mg/m3 

The converted value is then adjusted to an equivalent continuous exposure as follows: 

BMCLADJ = 272 mg/m3

 × 6 hours/24 hours × 5 days/7 days = 48.5 mg/m3 

To derive the RfC, several UFs are applied to the BMCLADJ. As discussed in Section 

5.1.3, a UF of 3 is used as an effect-level extrapolation factor, analogous to a LOAEL-toNOAEL UF, because the BMC is considered an adverse-effect level. In addition, the standard 

UFs of 3 for interspecies extrapolation for inhalation studies and of 10 for intraspecies variability 

are used. Also as discussed in Section 5.1.3, a UF of 3 for database deficiencies is used due to 

the absence of a two-generation reproductive/developmental toxicity study for benzene. Finally, 

a partial UF of 3 was used to extrapolate from subchronic to chronic exposure. This partial value 

was selected on the basis of the observation that hematologic fluctuations such as reductions in 

RBCs and WBCs in the high-dose mice were noted at interim sacrifice (14 days) as well as at 

termination (91 days), suggesting that the responses occurred early in the exposure cycle and 

then remained comparatively unchanged. Therefore, an overall UF of 3 × 3 × 10 × 3 × 3 = 1000 

is used in the calculation of the chronic inhalation RfC from the BMCLADJ, as follows: 

RfC = BMCLADJ ÷ UF = 48.5 mg/m3

 ÷ 1000 = 5 × 10-2 mg/m3 

This value is in good agreement with the RfC of 3 × 10-2 mg/m3 calculated from the BMC from 

the Rothman et al. (1996a) human study. 

Similarly, for comparison purposes, a chronic inhalation RfC can be derived from the 

NOAEL of 30 ppm observed for hematologic effects in the Ward et al. (1985) experimental 

animal study. First, the NOAEL is converted to mg/m3

 and adjusted to equivalent continuous 

exposure, as above: 

30 ppm × 78.11/24.45 = 95.8 mg/m3 

95.8 mg/m3 × (6 hours/24 hours) × 5 days/7 days = 17.1 mg/m3 

126
(ppm equivalence across species was assumed; this is identical to using EPA’s inhalation 

dosimetry methodology with RGDRr

=1 [U.S. EPA, 1994]). UFs are identical to those employed 

above, except that no NOAEL-to-LOAEL UF is used, thus the overall UF is 300, resulting in an 

RfC of 

NOAELADJ ÷ UF = 17.1 ÷ 300 = 6 × 10-2 mg/m3 

This value is also in good agreement with the RfC of 3 × 10-2 mg/m3 derived from the Rothman 

et al. (1996a) study. 

It should be noted, however, that other experimental animal studies have reported 

significant hematologic effects at benzene exposures of 10–25 ppm, which are lower than the 

NOAEL of 30 ppm from the Ward et al. (1985) study. These studies have insufficient data for 

dose-response modeling, and they used shorter exposure durations and/or fewer experimental 

animals than did the Ward et al. study; nonetheless, they observed statistically significant 

hematologic effects at 10–25 ppm. Baarson et al. (1984), for example, exposed male C57BL/6J 

mice (five/group) to 10 ppm benzene 6 hours/day, 5 days/week for 178 days and observed 

statistically significant reductions in blood lymphocytes at each of the three monitoring time 

points (32, 66, and 178 days) as compared with controls. The magnitude of the reduction in 

lymphocytes ranged from about 53% at 32 days to about 68% at 178 days. Cronkite et al. (1985) 

exposed male and female C57BL/6 BNL mice to various concentrations of benzene 6 hours/day, 

5 days/week for 2 weeks and observed no decrease in blood lymphocytes at 10 ppm; however, 

the investigators did observe a statistically significant reduction of about 21% at 25 ppm as 

compared with controls (5–10 mice/group). Thus, lower RfCs than those calculated above for 

the Ward et al. (1985) study are possible, based on other experimental animal results. In the 

most extreme case, using a LOAEL of 10 ppm and an overall UF of 3000 yields a LOAELADJ of 

5.7 mg/m3

 and an RfC of 2 × 10-3 mg/m3. 

5.2. ORAL REFERENCE DOSE (RfD) 

5.2.1. Choice of Principal Study and Critical Effect 

As with the inhalation RfC (Section 5.1), the human occupational inhalation study of 

Rothman et al. (1996a) was selected as the principal study, and reduced ALC was selected as the 

critical effect, for the derivation of the chronic oral RfD. This study was selected because it was 

a well-conducted human exposure study that demonstrated a dose-response relationship for 

hematologic effects, responses that are considered to be among the more sensitive indices of 

benzene toxicity (see Section 5.1.1 for more details). Furthermore, no relevant human data are 

available to evaluate hematologic effects following oral exposure. For comparison purposes, a 

chronic oral RfD is also derived from the chronic experimental animal gavage study of the NTP 

(1986) on the basis of the same critical effect of reductions in lymphocyte count (see Section 

5.2.5). 

5.2.2. Conversion of Inhalation Exposure to Equivalent Oral Dose Rate 

As discussed in Section 5.1.2, BMD modeling of the ALC data of Rothman et al. (1996a) 

yielded a BMC of 13.7 ppm (8-hour TWA) and a BMCL of 7.2 ppm (8-hour TWA) for the 

127
default BMR of one standard deviation change from the control mean response. This default 

definition of a BMR for continuous endpoints corresponds to an excess risk of approximately 

10% for the proportion of individuals below the 2nd percentile (or above the 98th percentile) of 

the control distribution for normally distributed effects (see U.S. EPA, 2000b). As suggested in 

the draft Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b), the BMCL is 

chosen as the point of departure for the RfD derivation. Calculation of an equivalent oral dose 

rate from the inhalation BMCL of 7.2 ppm (8-hour TWA) is shown below. 

The BMCL is first converted to mg/m3

 using the molecular weight of 78.11 for benzene 

and assuming 25°C and 760 mm Hg: 

BMCL (mg/m3) = 7.2 ppm × 78.11/24.45 = 23.0 mg/m3 

The converted value is then adjusted from the 8-hour TWA to an exposure concentration 

adjusted for continuous exposure using the default occupational minute volume (U.S. EPA, 

1994). 

BMCLADJ = BMCL (mg/m3) × (VEho/VEh) × 5 days/7 days 

BMCLADJ = 23.0 mg/m3 × (10 m3/20 m3) × 5 days/7 days = 8.2 mg/m3 

where: 

BMCLADJ = the BMCL dosimetrically adjusted to account for continuous exposure 

BMCL = occupational exposure level (8-hour TWA) 

VEho = human occupational default minute volume (10 m3

/8 hours) 

VEh = human ambient default minute volume (20 m3

/24 hours) 

In the support document for the benzene cancer assessment on IRIS (U.S. EPA, 1999a), 

EPA provided a simple method for extrapolating benzene-induced cancer risk from the 

inhalation to oral route. The same method is applied here for noncancer (hematopoietic) effects. 

The method is based on the relative efficiency of benzene absorption across routes of exposure, 

especially pulmonary and gastrointestinal barriers. An inhalation absorption rate of 50% and an 

oral absorption rate of 100% were used to calculate the absorbed benzene dose. These values are 

based on human inhalation absorption studies (Nomiyama and Nomiyama, 1974; Pekari et al., 

1992; Srbova et al., 1950) and the study of Sabourin et al. (1987), which compared inhalation 

and oral absorption in rats and mice. 

Sabourin et al. (1987) found that the retention of 14C-benzene by rats and mice during a 

6-hour exposure decreased as exposure concentration increased. Retention decreased from 33 ± 

6% to 15 ± 9% for rats and from 50 ± 1% to 10 ± 2% for mice as exposure concentration 

increased from 32 to 3,200 mg/m3

 (10 to 1,000 ppm). In the same study, gastrointestinal 

absorption of benzene administered by gavage was > 97% for doses between 0.5 and 150 mg/kg 

body weight. At oral doses below 15 mg/kg, > 90% of the 14C excreted was in the urine as 

128
nonethyl acetate-extractable material. At higher doses, an increasing percentage of the orally 

administered benzene was exhaled unmetabolized. Thus, in the dose range represented by the 

BMCL from the study by Rothman et al. (a, b), absorption of a comparable oral dose was 

assumed to be 100%. See also U.S. EPA (1999a) for more details about the route-to-route 

extrapolation of benzene inhalation results to oral exposures. 

To calculate an equivalent oral dose rate, the BMCLADJ is multiplied by the default 

inhalation rate, multiplied by 0.5 to correct for the higher oral absorption, and divided by the 

standard default human body weight of 70 kg (U.S. EPA, 1988): 

8.2 mg/m3

 × 20 m3/day × 0.5 ÷ 70 kg = 1.2 mg/kg/day 

5.2.3. RfD Derivation 

To calculate an RfD using the BMCLADJ-equivalent oral dose rate value of 1.2 

mg/kg/day, several UFs were applied. First, because the BMC is considered to be an adverse 

effect level, an effect-level extrapolation factor analogous to the LOAEL-to-NOAEL UF is used. 

EPA is planning to develop guidance for applying an effect level extrapolation factor to a BMD. 

A factor of 3 will be used in this analysis, as in Section 5.1.3. For a more serious effect, a larger 

factor, such as 10, might be selected. Second, a factor of 10 was used for intraspecies 

differences in response (human variability) as a means of protecting potentially sensitive human 

subpopulations. Third, a partial UF of 3 was used for subchronic-to-chronic extrapolation, as 

discussed in Section 5.1.3. Finally, a UF of 3 was chosen to account for database deficiencies 

due to the lack of a two-generation reproductive/developmental toxicity study for benzene. 

Therefore, an overall UF of 3 x 10 x 3 x 3 = 300 is used to calculate the chronic oral RfD, as 

follows: 

RfD = equivalent oral dose rate ÷ UF = 1.2 mg/kg/day ÷ 300 = 4 × 10-3 mg/kg/day. 

Use of a modifying factor of 3 to account for uncertainty in the route-to-route extrapolation was 

considered; however, it was deemed unnecessary. The RfD is based on human data for a 

sensitive endpoint; thus, it was felt that the composite UF of 300 provides sufficient protection. 

129
5.2.4. Comparison Analysis Based on the LOAEL 

A median 8-hour TWA concentration of 7.6 ppm was identified from the Rothman et al. 

(1996a) study as a LOAEL for the critical effect of reduced lymphocyte counts. An equivalent 

oral dose rate can be calculated from the LOAEL, as shown in Section 5.2.2 above and in 

abbreviated form below. First, a LOAELADJ is calculated by converting to mg/m3 and adjusting 

from the occupational ventilation rate and intermittent work-week schedule to a continuous 24-

hour exposure, 7 days/week: 

LOAEL (mg/m3) = 7.6 ppm × 78.11/24.45 = 24.3 mg/m3 

LOAELADJ = 24.3 mg/m3 × (10 m3/20 m3) × 5 days/7 days = 8.7 mg/m3 

Next, an inhalation absorption rate of 50% was used to calculate the absorbed benzene 

dose, and oral absorption was assumed to be 100%, as discussed in Section 5.2.2. To calculate 

an equivalent oral dose rate, the LOAELADJ is multiplied by the default inhalation rate, 

multiplied by 0.5 to correct for the higher oral absorption, and divided by the standard default 

human body weight of 70 kg (U.S. EPA, 1988): 

8.7 mg/m3

 × 20 m3/day × 0.5 ÷ 70 kg = 1.2 mg/kg/day 

To calculate an RfD using the LOAELADJ-equivalent oral dose rate value of 1.2 

mg/kg/day, the following UFs were selected: a factor of 10 to account for using a LOAEL due 

to the lack of an appropriate NOAEL, a factor of 10 for intraspecies differences in response 

(human variability), a factor of 3 for subchronic-to-chronic extrapolation, and a factor of 3 for 

database deficiencies, as above. Therefore, an overall UF of 10 × 10 × 3 × 3 = 1000 has been 

used to calculate the chronic oral RfD, as follows: 

RfD = equivalent oral dose rate ÷ UF = 1.2 mg/kg/day ÷ 1000 = 1 × 10-3 mg/kg/day. 

This value is in good agreement with the RfD of 4 × 10-3 mg/kg/day calculated from the BMC. 

5.2.5. Comparison Analysis Based on the NTP (1986) Experimental Animal Study 

The proposed RfD is based on the Rothman et al. (1996a) study because it was decided 

that it was preferable to use the human data rather than the experimental animal data. The NTP 

(1986) study rather than the Hsieh et al. (1988b) study was used for the comparison analysis 

because it was much larger (50 vs. 5 animals/dose group), it was conducted for 2 years (vs. 28 

days) with blood measurements made at multiple time points, and it examined both male and 

female rats and mice (vs. just male mice). The Hsieh et al. (1988b) study provided a lower 

LOAEL (8 mg/kg/day vs. 18 mg/kg/day adjusted to 7 days/week); however, the lower LOAEL is 

not crucial, because the main analysis currently uses a BMD approach. Both studies yielded 

similar BMDLs for decreased ALC (0.7 mg/kg/day for the male rat NTP data vs. 1.4 mg/kg/day 

for the Hsieh et al. data; both based on a linear model with transformed doses and a BMR of one 

standard deviation change from the control mean). 

130
A chronic oral RfD was also derived from the NTP (1986) experimental animal study for 

comparison with the RfD of 1 × 10-3 mg/kg/day based on the Rothman et al. (1996a) human 

study. In the NTP (1986) study, F344 rats and B6C3F1 mice of both sexes were administered 

benzene by gavage 5 days/week for 103 weeks (see also Section 4.2.1.2). Male rats (50/group) 

were administered doses of 0, 50, 100, or 200 mg/kg and females (50/group) were administered 

doses of 0, 25, 50, or 100 mg/kg. B6C3F1 mice (50/sex/group) were administered doses of 0, 

25, 50, or 100 mg/kg. Blood was drawn from 10 randomly preselected animals per 

species/sex/dose group at 12, 15, 18, and 21 months and from all animals at the terminal kill at 

24 months. Additional groups of 10 animals of each sex and species were administered benzene 

for 51 weeks at the same doses of the 2-year study, and blood was drawn at 0, 3, 6, 9, and 12 

months. This study identified a LOAEL of 25 mg/kg for leukopenia and lymphocytopenia in 

female F344 rats and male and female B6C3F1 mice and of 50 mg/kg in male F344 rats. These 

were the lowest doses tested and, thus, no NOAEL was identified. 

Reduction in lymphocyte count was selected as the critical effect, and attempts were 

made to model the dose-response relationships using a BMD modeling approach. However, the 

dose-response analysis of the NTP data was problematic for a number of reasons. First, only the 

grouped data, not the individual animal data, were provided. Second, as discussed in Appendix 

N of the NTP (1986) report, not all of the groups were comparable for a variety of reasons, 

including experiment design, differences in bleeding methods and time of bleeding, and 

differences in measurement instrumentation. Third, the grouped data exhibited substantial 

variability, and baseline measurements (vehicle control groups) were not constant over time. 

Furthermore, for the rats, the strong dose effect in lymphocyte count was accompanied by 

significant temporal variability, which appears to be largely due to temporal variability in the 

vehicle control group means. For these reasons, it was not possible to do a pooled analysis 

across time points. Instead, modeling was performed for each dataset in two data groupings 

within which the datasets are comparable (months 6 and 9; and months 12, 15, 18, and 21), and 

ranges of results are presented. Each of these datasets had at most 10 animals/dose, so the doseresponse results are not very robust. The males of both species exhibited more dramatic and 

consistent reductions in lymphocyte count than did the females, but it was not clear a priori 

which species was more sensitive; therefore, dose-response analyses were performed for both the 

male mouse (NTP Table N8) and the male rat (NTP Table N4). 

The continuous linear, polynomial, and power models in EPA’s Benchmark Dose 

Modeling Software (version 1.20) were used for the modeling. The software estimates the 

parameters using the method of maximum likelihood. Most of the data were supralinear (i.e., the 

magnitude of the reductions in lymphocyte count decreased with increasing unit dose) and not 

amenable to modeling with restricted models (in this case, with parameters constrained to be 

nonpositive), as suggested in EPA’s draft Benchmark Dose Technical Guidance Document (U.S. 

EPA, 2000b). Therefore, two different strategies were attempted. The first was to use 

unrestricted polynomial models and exclude the high-dose group when necessary to fit the 

models, and the second was to transform the dose data before modeling. Dropping the high-dose 

groups alone did not resolve the problem because there was still too much supralinear curvature 

in the two lower dose groups to fit restricted models. Results are summarized below in Tables 

11 and 12, for untransformed and transformed data, respectively. 

131
For each dataset, the selected model was chosen on the basis of the lowest Akaike’s 

Information Criterion value with consideration of the graphical display, as suggested in the draft 

BMD technical guidance document. For selecting between models within a family of models 

(e.g., between a linear and a two-degree polynomial model), consideration was given to the loglikelihood values to evaluate the statistical significance of adding an extra parameter. In some 

cases there was no model that adequately fit the data. The draft technical guidance document 

suggests a p-value of 0.1 for model fit. In addition, there was substantial variability in these 

data, but the variability appeared to be random and not amenable to modeling. Therefore, 

constant variance was assumed for all the models, although in some cases the variances failed 

the test for homogeneity. 

In the absence of a clear definition for an adverse effect for this endpoint, a default BMR 

of one standard deviation change from the control mean response was selected as suggested in 

the draft BMD technical guidance document. This definition of the BMR is highly sensitive to 

the substantial variability in data such as these; thus, the BMR itself is not very robust. The 

usefulness of this default definition would be strengthened by the use of a larger dataset of 

historical control data, but such data were not located. The software uses the estimated 

“constant” standard deviation as the standard deviation for all the group means. The 95% lower 

confidence limit (BMDL) on the resulting BMD was then calculated using the likelihood profile 

method. 

Because many of the datasets could not be modeled using the available models without 

dropping the parameter constraints and sometimes even excluding the highest dose group (Table 

11), a second modeling approach was tried in which the dose levels were transformed before the 

dose-response modeling was conducted, as was done for the human data modeled in Section 

5.1.2. The dose levels were transformed according to the formula: transformed dose = 

ln(dose + 1). Results of the dose-response modeling using the transformed doses are presented 

in Table 12. 

The results in Table 12 suggest that the male rat is more sensitive than the male mouse to 

lymphocyte count reductions from exposure to benzene in this NTP (1986) gavage bioassay, 

because the ranges of BMDs/BMDLs are substantially lower for the male rat, especially for year 

2. The ranges for the male rat are also fairly tight, and the models selected provide good fits to 

all the male rat datasets. However, all but one of the calculated BMDs for the male rat are over 

an order of magnitude below the lowest exposure dose of 50 mg/kg. Ideally, BMDs should be 

closer to the low end of the range of observation (i.e., the range of the actual exposure doses) to 

reduce the impacts of model selection and the uncertainties inherent in extrapolating to lower 

doses. 

Nevertheless, data from two drinking water studies (Hsieh et al., 1988b; White et al., 

1984) provide support for selecting a BMD in this range. These two studies were of shorter 

duration and used fewer experimental animals than did the NTP (1986) study; however, they do 

provide dose-response data for BMD modeling and they also have the advantage of being 

drinking water studies; thus the benzene exposure scenario is more relevant to human oral 

benzene exposures. In one study, Hsieh et al. (1988b) exposed male CD-1 mice (five/group) to 

0, 8, 40, or 180 mg/kg/day benzene in drinking water for 28 days. Hematologic effects were 

132
Table 11. BMD modeling results of the NTP (1986) male mouse and male rat 

lymphocyte counts, with untransformed data 

Dataset Model 

Variance 

homogeneity Fit 

BMDa 

(mg/kg) 

BMDLa 

(mg/kg) 

Male mouse 

6-month unrestricted 2-degree 

polynomial 

OK yes 

p=0.13 

20.52 12.96 

9-month unrestricted 2-degree 

polynomial 

no yes 

p=0.72 

23.06 13.70 

year 1 range 20.52–23.06 12.96–13.70 

12-month unrestricted 2-degree 

polynomial 

OK nob 

p=0.005 

16.68b 11.41b 

15-month linear no yes, p=0.98 51.94 36.80 

18-month linear no yes, p=0.16 35.50 26.92 

21-month linear no yes, p=0.25 42.91 30.44 

year 2 range 35.50–51.94 26.92–36.80 

Male rat 

6-month unrestricted 2-degree 

polynomial 

OK yes 

p=0.56 

15.70 11.79 

9-month unrestricted 2-degree 

polynomial 

no yes 

p=0.37 

16.88 12.72 

year 1 range 15.70–16.88 11.79–1272 

12-month; w/o 

high-dose 

group 

unrestricted 2-degree 

polynomial 

OK yesc 10.51 7.3 

15-month; w/o 

high-dose 

group 

unrestricted 2-degree 

polynomial 

OK yesc 7.56 5.56 

18-month; w/o 

high-dose 

group 

unrestricted 2-degree 

polynomial 

no yesc 13.84 9.07 

21-month; w/o 

high-dose 

group 

unrestricted 2-degree 

polynomial 

OK yesc 7.21 5.03 

year 2 range 7.21–13.84 5.03–9.07 

a

Unadjusted animal doses in mg/kg. b

These data could not be fit, and the BMD and BMDL are excluded from the reported ranges. c

Models were judged to fit on the basis of the graphical displays and 32-residuals; however, there were insufficient

degrees of freedom for calculation of a p-value.

133
Table 12. BMD modeling results of the NTP (1986) male mouse and male rat 

lymphocyte counts, with transformed dose data 

Dataset Model 

Variance 

homogeneity Fit 

BMDa 

(mg/kg) 

BMDLa 

(mg/kg) 

Male Mouse 

6-month 2-degree polynomial OK borderline p=0.047 19.68 6.57 

9-month linear no yes, p=0.35 9.07 4.05 

year 1 range 9.07–19.68 4.05–6.57 

12-month linear OK yes, p=0.30 3.74 2.32 

15-month power no yes, p=0.31 47.46 18.55 

18-month power no borderline p=0.09 28.93 13.99 

21-month power no yes, p=0.15 23.34 5.80 

year 2 range 3.74–47.46 2.32–18.55 

Male rat 

6-month power OK yes, p=0.30 9.92 4.52 

9-month linear no yes, p=0.11 3.71 2.30 

year 1 range 3.71–9.92 2.30–4.52 

12-month linear no yes, p=0.22 1.34 0.95 

15-month linear OK yes, p=0.93 1.34 0.95 

18-month linear no yes, p=0.22 2.73 1.74 

21-month linear OK yes, p=0.54 1.69 1.10 

year 2 range 1.34–2.73 0.95–1.74 

a

Unadjusted animal dose in mg/kg, after transforming the results back according to the formula: dose = 

exp(transformed dose)  1. 

observed at all exposure levels. BMD modeling of the ALC yielded a BMD of 2.2 mg/kg/day 

and a BMDL of 1.4 mg/kg/day, based on a linear model with transformed doses and a BMR of 

one standard deviation change from the control mean, as above. In the second study, White et al. 

(1984) exposed female B6C3F1 mice to 0, 12, 195, or 350 mg/kg/day benzene in drinking water 

for 30 days. BMD modeling of the ALC (five to six mice/group) resulted in a BMD of 11.6 

mg/kg/day and a BMDL of 5.3 mg/kg/day (also based on a linear model with transformed doses 

and a BMR of one standard deviation change from the control mean, as above). 

The results in Table 12 from BMD modeling of the male rat ALC data from the NTP 

(1986) study show the lowest BMDL of about 1 mg/kg at three time points in the second year; 

therefore, this was selected as the point of departure for an RfD calculation. Adjusting for 

134
exposure 7 days/week yields a BMDLADJ of 0.7 mg/kg/day. As discussed in previous 

subsections, a UF of 3 is used as an effect-level extrapolation factor, analogous to a LOAEL-toNOAEL UF, because the BMD is considered to be an adverse-effect level. In addition, the 

standard UFs of 10 for interspecies extrapolation for oral studies and for intraspecies variability 

are used. Finally, a UF of 3 for database deficiencies is used, as in the previous derivations. 

These values yield an overall UF of 1000 and an RfD as follows: 

RfD = BMDLADJ ÷ UF = 0.7 mg/kg/day ÷ 1000 = 7 × 10-4 mg/kg/day 

This RfD value is in reasonably good agreement (within an order of magnitude) with the RfD 

value of 4 × 10-3 mg/kg/day derived from the Rothman et al. (1996a) human inhalation study. 

For comparison purposes, a chronic oral RfD can be derived from the LOAEL of 25 

mg/kg identified for hematologic effects in the NTP (1986) study (there was no NOAEL). 

Adjusting from 5-day to 7-day exposure yields a LOAELADJ of 18 mg/kg/day, which can be used 

to calculate an RfD for benzene as follows: 

RfD = LOAELADJ ÷ UF = 18 mg/kg/day ÷ 3000 = 6 × 10-3 mg/kg/day, 

where the combined UF of 3000 is made up of component factors of 10 for LOAEL-to-NOAEL 

extrapolation, 10 for interspecies extrapolation, 10 for intraspecies variability, and 3 for database 

deficiencies. This value is in good agreement with the RfD of 4 × 10-3 mg/kg/day calculated 

from the BMD analysis of the Rothman et al. (1996a) human data. 

5.3. DOSE-RESPONSE SUMMARY 

For the derivation of a chronic inhalation RfC, the human occupational inhalation study 

of Rothman et al. (1996a) was selected as the principal study, and the inhalation study of Ward 

et al. (1985) was selected as a supporting experimental animal study. These two studies 

provided the best dose-response data for quantitatively evaluating the potential human health 

risks due to inhalation of benzene. The cross-sectional study of Rothman et al. (1996a) was 

designated as the principal study because it is a human exposure study with well-quantified 

benzene exposure estimates and data on sensitive hematological endpoints. The study compared 

the hematologic evaluations of 44 workers occupationally exposed to benzene with those of 44 

age- and gender-matched unexposed controls, all based in Shanghai, China. ALC, WBCs, 

RBCs, and platelets were all significantly decreased, whereas MCV was significantly increased 

in the group of exposed workers (median 8-hour TWA of 31 ppm [99 mg/m3

]), compared with 

the age- and sex-matched control group. In the low-dose subjects (< 31 ppm, median 8-hour 

TWA of 13.6 ppm [43.4 mg/m3

]), the ALC, RBCs, and platelet count were reduced as compared 

with controls. Similarly, in a selected subgroup exposed to a median 8-hour TWA of 7.6 ppm 

(23 mg/m3) benzene, a statistically significant difference in ALC versus controls was observed 

(p<0.03). Furthermore, a dose-response relationship was established between the ALC and 

benzene exposure, as monitored by organic vapor passive dosimetry and the level of benzene 

metabolites in the urine. The median 8-hour TWA of 7.6 ppm was designated the LOAEL for 

these effects and was used to calculate a chronic inhalation RfC for benzene hematotoxicity in 

humans for comparison with the RfC derived in the primary analysis using BMD modeling of 

the ALC exposure-response data. 

135
BMD modeling of the ALC exposure-response data was conducted using the continuous 

models from EPA’s Benchmark Dose Modeling Software (version 1.20). In order to fit the data 

with one of the available models, the data were first transformed according to the equation 

d1 = ln(d + 1). The exposure-response data were then fit using the continuous linear model, 

which provided a good fit. The two-degree polynomial model and the power model also 

provided adequate fits to the data; however, the linear model was selected because it was more 

parsimonious and the additional parameters were not statistically significant. In the absence of a 

clear definition for an adverse effect for this continuous endpoint, a default BMR of one standard 

deviation change from the control mean response was selected, as suggested in EPA’s draft 

Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b). This default definition of a 

BMR for continuous endpoints corresponds to an excess risk of approximately 10% for the 

proportion of individuals below the 2nd percentile (or above the 98th percentile) of the control 

distribution for normally distributed effects. Transforming the resulting BMC and BMCL back 

to the original exposure scale yielded a BMC of 13.7 ppm (8-hour TWA) and a BMCL of 7.2 

ppm (8-hour TWA). 

As suggested in the draft BMD technical guidance document, the BMCL was chosen as 

the point of departure for the RfC derivation. After converting to mg/m3

 and adjusting for 

continuous exposure, a BMCLADJ of 8.2 mg/m3

 was obtained. Dividing this value by an overall 

UF of 300 yields a chronic inhalation RfC of 3 × 10-2 mg/m3

, based on BMD modeling of the 

ALC data from the Rothman et al. (1996a) human study. Because the BMC is considered to be 

an adverse-effect level, an effect-level extrapolation factor analogous to the LOAEL-to-NOAEL 

UF was used. EPA is planning to develop guidance for applying an effect-level extrapolation 

factor to a BMD. In the interim, a factor of 3 was used in this analysis. For a more serious 

effect, a larger factor, such as 10, might be selected. Additional factors of 10 for intraspecies 

variability, 3 for subchronic-to-chronic extrapolation (exposure range 0.7 to 16 years), and 3 for 

database deficiencies, due to the absence of a two-generation reproductive/developmental 

toxicity study for benzene, comprise the remainder of the 300 composite UF. These UFs, as well 

as the UFs for the comparison analyses and the RfD calculations, are summarized in Table 13. 

For comparison, an RfC was also calculated from the LOAEL of 7.6 ppm (8-hour TWA) 

from Rothman et al. (1996a). After converting to mg/m3

 and adjusting for continuous exposure, 

a LOAELADJ of 8.7 mg/m3

 was obtained. Dividing this value by an overall UF of 1000 yields an 

RfC of 9 × 10-3 mg/m3. The UF of 1000 was based on a factor of 10 to account for the use of a 

LOAEL because of the lack of an appropriate NOAEL, a factor of 10 for intraspecies differences 

in response (human variability), a partial UF of 3 for subchronic-to-chronic extrapolation, and a 

factor of 3 for database deficiencies in the absence of a two-generation reproductive/ 

developmental toxicity study for benzene. This result of 9 × 10-3 mg/m3

 based on the LOAEL is 

in good agreement with the result of 3 × 10-2 mg/m3 based on the BMCL. The BMD modeling 

approach is chosen as the primary analysis in this document because BMD modeling is a 

generally superior methodology that addresses some of the limitations of the LOAEL/NOAEL 

approach. For example, BMD modeling makes use of all the dose-response data, and the BMD 

is not restricted to being one of the doses used in the study. The BMD also provides a more 

136
Table 13. Summary of uncertainty factors used for deriving the RfC and RfDa 

Basis 

BMCL or BMDL (NOAEL or LOAEL)b 

Effect level 

extrapolation 

factor 

(LOAEL-toNOAEL UF) Interspecies UF Intraspecies UF Subchronic- to-chronic UF 

Database 

deficiencies 

UF 

Composite 

UF RfD or RfC 

RfC based on Rothman et al. (1996a) (human) 

8.2 mg/m3 

(8.7 LOAEL) 

3 

(10) 

NA 10 3 3 300 

(1000) 

0.03 mg/m3 

(0.009) 

Comparison RfC based on Ward et al. (1985) (rodent) 

48.5 mg/m3 

(17 NOAEL) 

3 

(NA) 

3 10 3 3 1000 

(300) 

0.05 mg/m3 

(0.06) 

RfD based on Rothman et al. (1996a) (human) 

1.2 

mg/kg/dayc 

(1.2 LOAEL) 

3 

(10) 

NA 10 3 3 300 

(1000) 

4 × 10-3 

mg/kg/day 

(1 × 10-3) 

Comparison RfD based on NTP (1986) (rodent) 

0.7 

mg/kg/day 

(18 LOAEL) 

3 

(10) 

10 10 NA 3 1000 

(3000) 

7 × 10-4 

mg/kg/day 

(6 × 10-3) 

aPrimary results are based on BMD analysis; comparison results using NOAEL/LOAEL approach are presented in parentheses where different. bThe BMC was based on a BMR of one standard deviation change from the control mean. All values are adjusted for continuous exposure. cOral BMDL (and LOAEL) for Rothman et al. (1996a) inhalation study was derived by route-to-route extrapolation with the assumptions that inhalation 

absorption was 50% and oral absorption was 100% in the dose range near the BMC. 

NA = not applicable. 

137
standardized point of comparison across endpoints and studies because it corresponds to a 

specific response level rather than being merely one of the doses used in a particular study. 

Support for this chronic inhalation RfC was provided by the experimental animal study of 

Ward et al. (1985). The subchronic inhalation study of Ward et al. (1985) was selected as a 

supporting study because it used a relatively long inhalation exposure duration and an adequate 

number of animals; it also provided dose-response data. The study identified both a LOAEL of 

300 ppm and a NOAEL of 30 ppm. Ward et al. exposed male and female CD-1 mice and 

Sprague-Dawley rats to 0, 1, 10, 30 or 300 ppm (0, 3.2, 32, 96 or 960 mg/m3

) benzene 6 

hours/day, 5 days/week for 91 days and measured various hematologic endpoints. The male 

mouse appears to be the most sensitive sex/species in this study. The exposure-response 

relationships for the different hematologic endpoints in male mice were modeled using a BMD 

modeling approach, and decreased HCT was chosen as the critical effect. 

BMD modeling was conducted using the continuous models from EPA’s Benchmark 

Dose Modeling Software (version 1.20). The linear, polynomial, and power models all resulted 

in the same BMC and BMCL estimates; however, the linear model had better results for the fit 

statistics. Thus, the continuous linear model was selected. In the absence of a clear definition 

for an adverse effect for this continuous endpoint, a default BMR of one standard deviation for 

an adverse effect for this continuous endpoint, a default BMR of one standard deviation change 

from the control mean response was selected, as suggested in EPA’s draft Benchmark Dose 

Technical Guidance Document (U.S. EPA, 2000b). A BMC of 100.7 ppm and a BMCL of 85.0 

ppm were obtained. It should be noted that because of the large dose spacing between the 30 

ppm and the 300 ppm groups, there is some uncertainty about the actual shape of the doseresponse curve in the region of the BMC and, thus, some corresponding uncertainty about the 

values of the BMC and BMCL estimates. 

ALCs were not reported in Ward et al. (1985), so this endpoint could not be compared 

with the human ALC results. Total WBC counts were reported and exhibited the largest percent 

change in response between the control and the 300 ppm group; however, the data for this 

endpoint also had substantial variance. In addition, because the BMR used for this analysis is a 

function of the standard deviation, WBC counts did not yield the lowest BMC estimate. The 

actual lowest BMC estimates were obtained for increased mean cell Hgb (78 ppm; BMCL = 67 

ppm) and increased mean cell volume (79 ppm; BMCL = 68 ppm); however, these endpoints are 

probably not adverse per se. On the other hand, they are likely to be compensatory effects—and 

thus markers of toxicity—and one could probably justify using them as the critical effects. In 

any event, the BMC estimates are not much different from the BMC of 100 ppm obtained for 

decreased HCT. The results are also similar for total blood Hgb (BMC = 104 ppm; BMCL = 88 

ppm). RBC results were in between those for MCV and MCH and those for HCT and total Hgb; 

however, the model fits were not adequate for the RBC data. Thus, the RBC results have more 

uncertainty. 

As suggested in the draft BMD technical guidance document, the BMCL was chosen as 

the point of departure for the RfC derivation. After converting to mg/m3

 and adjusting for 

continuous exposure, a BMCLADJ of 48.5 mg/m3

 was obtained for the critical effect of decreased 

HCT. To derive the RfC, several UFs are applied to the BMCLADJ. As discussed above, a UF of 

3 is used as an effect-level extrapolation factor, analogous to a LOAEL-to-NOAEL UF, because 

138
the BMC is considered an adverse-effect level. In addition, the standard UFs of 3 for 

interspecies extrapolation for inhalation studies and of 10 for intraspecies variability are applied. 

A UF of 3 for database deficiencies is used due to the absence of a two-generation 

reproductive/developmental toxicity study for benzene. Finally, a partial UF of 3 was used to 

extrapolate from subchronic to chronic exposure. This partial value was selected on the basis of 

the observation that hematologic fluctuations such as reductions in RBCs and WBCs in the highdose mice were noted at interim sacrifice (14 days) as well as at termination (91 days), 

suggesting that the responses occurred early in the exposure cycle and then remained 

comparatively unchanged. Dividing the BMCLADJ by the overall UF of 1000 yields a chronic 

inhalation RfC of 5 × 10-2 mg/m3

. This value is in good agreement with the RfC of 3 × 10-2 

mg/m3, based on BMD modeling of the ALC data from the Rothman et al. (1996a) human study. 

Similarly, for comparison purposes, a chronic inhalation RfC can be derived from the 

NOAEL of 30 ppm observed for hematologic effects in the Ward et al. (1985) experimental 

animal study. First, the NOAEL is converted to mg/m3 and adjusted to equivalent continuous 

exposure, yielding 17.1 mg/m3

. UFs are identical to those employed above, except that no 

NOAEL-to-LOAEL UF is used; thus, the overall UF is 300. Dividing 17.1 by 300 results in an 

RfC of 6 × 10-2 mg/m3

. This value is also in good agreement with the RfC derived from the 

Rothman et al. (1996a) study. 

For derivation of an RfD, the human occupational inhalation study of Rothman et al. 

(1996a) was again selected as the principal study, and the 103-week gavage study conducted by 

the NTP (1986) in F344 rats and B6C3F1 mice was selected as a supporting experimental animal 

study. These studies provided the best dose-response data for quantitatively evaluating the 

potential human health risks due to oral benzene exposure. The cross-sectional study of 

Rothman et al. (1996a) is designated as the principal study because it is a human exposure study, 

with well-quantified exposure estimates and data on sensitive hematological endpoints. 

As with the RfC, the BMCL of 7.2 ppm (8-hour TWA) for the default BMR of one 

standard deviation change from the control mean response for the critical effect of reduced ALC 

in Rothman et al. (1996a) was used as the point of departure for the derivation of the RfD. After 

converting the units, correcting for continuous exposure, and adjusting for the route-to-route 

extrapolation from inhalation to oral exposure, a BMCLADJ-equivalent oral dose rate of 1.2 

mg/kg/day was obtained. The route-to-route extrapolation assumes 50% absorption of inhaled 

doses and 100% absorption of oral doses. These assumptions were based on experimental data 

(Nomiyama and Nomiyama, 1974; Pekari et al., 1992; Srbova et al., 1950; Sabourin et al., 1987) 

and are the same as those used by EPA for route-to-route extrapolation of the inhalation cancer 

risk estimates to obtain cancer risk estimates for oral exposures (U.S. EPA, 1999a). Dividing the 

BMCLADJ-oral of 1.2 mg/kg/day by an overall UF of 300 yields a chronic oral RfD of 4 × 10-3 

mg/kg/day, based on BMD modeling of the ALC data from the Rothman et al. (1996a) human 

study. Because the BMC is considered to be an adverse-effect level, an effect-level 

extrapolation factor analogous to the LOAEL-to-NOAEL UF was used. EPA is planning to 

develop guidance for applying an effect-level extrapolation factor to a BMD. In the interim, a 

factor of 3 was used in this analysis. Additional factors of 10 for intraspecies variability, 3 for 

subchronic-to-chronic extrapolation (exposure range 0.7 to 16 years), and 3 for database 

deficiencies due to the absence of a two-generation reproductive/developmental toxicity study 

for benzene comprise the remainder of the 300 composite UF. 

139
For comparison, an RfD was also calculated from the LOAEL of 7.6 ppm (8-hour TWA) 

from Rothman et al. (1996a). After unit conversion, correction for continuous exposure, routeto-route extrapolation, and division by a combined UF of 1000, an RfD of 1 × 10-3 mg/kg/day 

was derived from the LOAEL. The route-to-route extrapolation was based on an assumption of 

50% absorption of inhaled doses and 100% absorption of oral doses, as above. The combined 

UF of 1000 was based on a factor of 10 to account for using a LOAEL because of the lack of an 

appropriate NOAEL, a factor of 10 for intraspecies variability, a factor of 3 for subchronic-tochronic extrapolation, and a factor of 3 for database deficiencies. This RfD value of 1 × 10-3 

mg/kg/day is in good agreement with the value of 4 × 10-3 mg/kg/day calculated from the BMC. 

Support for this chronic oral RfD was provided by the experimental animal study of the 

NTP (1986). In this study, F344 rats and B6C3F1 mice of both sexes were administered benzene 

by gavage 5 days/week for 103 weeks (see also Section 4.2.1.2). For rats, males (50/group) were 

administered doses of 0, 50, 100, or 200 mg/kg, and females (50/group) were administered doses 

of 0, 25, 50, or 100 mg/kg. B6C3F1 mice (50/sex/group) were administered doses of 0, 25, 50, 

or 100 mg/kg. Blood was drawn from 10 randomly preselected animals per species/sex/dose 

group at 12, 15, 18, and 21 months and from all animals at the terminal kill at 24 months. 

Additional groups of 10 animals of each sex and species were administered benzene for 51 

weeks at the same doses of the 2-year study, and blood was drawn at 0, 3, 6, 9, and 12 months. 

This study identified a LOAEL of 25 mg/kg for leukopenia and lymphocytopenia in female F344 

rats and male and female B6C3F1 mice and 50 mg/kg in male F344 rats. These were the lowest 

doses tested and, thus, no NOAEL was identified. 

Reduction in lymphocyte count was selected as the critical effect, and attempts were 

made to model the dose-response relationships using a BMD modeling approach. The males of 

both species exhibited more dramatic and consistent reductions in lymphocyte count than did the 

females, but it was not clear a priori which species was more sensitive; therefore, dose-response 

analyses were performed on datasets for various time points for both the male mouse (NTP Table 

N8) and the male rat (NTP Table N4). Various continuous models were used to fit the different 

datasets, as appropriate, and one standard deviation change from the control mean was used as 

the BMR, as discussed above. The modeling results suggested that the male rat is more sensitive 

than the male mouse to lymphocyte count reductions from exposure to benzene in this NTP 

gavage bioassay. However, all but one of the calculated BMDs for the male rat were over an 

order of magnitude below the lowest exposure dose of 50 mg/kg. Ideally, BMDs should be 

closer to the low end of the range of observation (i.e., the range of the actual exposure doses) to 

reduce the impacts of model selection and the uncertainties inherent in extrapolating to lower 

doses. Nonetheless, BMD modeling of two subchronic drinking water studies (Hsieh et al., 

1988b; White et al., 1984) supported the selection of a BMD in the range of BMDs estimated 

from the NTP (1986) rat data, and a BMDL of 1 mg/kg was selected as the point of departure. 

Adjusting the BMDL of 1 mg/kg for exposure 7 days/week yielded a BMDLADJ of 0.7 

mg/kg/day. A UF of 3 was used as an effect-level extrapolation factor, analogous to a LOAELto-NOAEL UF, because the BMD is considered to be an adverse-effect level. Also, UFs of 10 

for interspecies extrapolation for oral studies, 10 for intraspecies variability, and 3 for database 

deficiencies due to the absence of a two-generation reproductive/developmental study were 

applied, resulting in a composite UF of 1000. Dividing the BMDLADJ by 1000 yields an RfD of 

7 × 10-4 mg/kg/day. This RfD value is in reasonably good agreement (within an order of 

140
magnitude) with the RfD value of 4 × 10-3 mg/kg/day derived from the Rothman et al. (1996a) 

human inhalation study. 

For comparison purposes, a chronic oral RfD was also calculated using the LOAEL from 

the NTP (1986) study. The LOAEL of 25 mg/kg was adjusted to a continuous exposure level of 

17.9 mg/kg/day and then divided by a UF of 3000 to derive an RfD of 6 × 10-3 mg/kg/day. The 

combined UF of 3000 is based on a factor of 10 for the absence of a NOAEL, a factor of 10 for 

interspecies extrapolation, a factor of 10 for intraspecies differences in response (human 

variability), and a factor of 3 for database deficiencies, as above. This value of 6 × 10-3 

mg/kg/day is in good agreement with the value 4 × 103 mg/kg/day derived from the Rothman et 

al. (1996a) human study. 

In summary, the chronic inhalation RfC values calculated on the basis of the human and 

the experimental animal data and using BMD modeling and NOAEL/LOAEL approaches are in 

good agreement, yielding values that range from 9 × 10-3 to 6 × 10-2 mg/m3

 (Table 14). This 

consistency in results provides increased confidence in the selected chronic inhalation RfC of 3 

× 10-2 mg/m3

, which is based on BMD modeling of the Rothman et al. (1996a) human data. 

Similarly, the chronic oral RfD values calculated on the basis of the human and the 

experimental animal data and using BMD modeling and NOAEL/LOAEL approaches are in 

generally good agreement, yielding values that range from 7 × 10-4 to 6 × 10-3 mg/kg/day. This 

consistency in results provides increased confidence in the selected chronic oral RfD of 4 × 10-3 

mg/kg/day, which is based on BMD modeling of the Rothman et al. (1996a) human data. 

The RfC and RfD values based on the Rothman et al. (1996a) human study were selected 

over the values derived from experimental animal studies because they are based on goodquality human data and therefore are not subject to the uncertainties inherent in interspecies 

extrapolation. Reference values based on BMD modeling were selected over those calculated 

using the NOAEL/LOAEL approach because BMD modeling is a superior methodology that 

makes better use of the exposure-response data. In any event, especially in the case of the RfC, 

all the estimates are in sufficiently good agreement as to be effectively indistinguishable. The 

overall confidence in this RfC and RfD assessment is medium. 

6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF 

HAZARD AND DOSE RESPONSE 

6.1. HUMAN HAZARD POTENTIAL 

Benzene is widely used as an industrial solvent, as an intermediate in chemical syntheses, 

and as a gasoline additive (NTP, 1994). Because of its widespread use, the potential for human 

exposure is great. The toxicity of benzene has been recognized for more than a century, and the 

141
Table 14. Summary of RfC and RfD estimates using human and experimental 

animal data, as well as BMD modeling and LOAEL/NOAEL approachesa 

Approach 

RfC (mg/m3) RfD (mg/kg/day) 

Human Rodent Human Rodent 

BMD modeling 3 × 10-2 5 × 10-2 4 × 10-3 7 × 10-4 

LOAEL/NOAE 

L 

9 × 10-3 6 × 10-2 1 × 10-3 6 × 10-3 

a

Selected values are in bold. 

biological impacts of benzene exposure have been extensively studied in humans and in 

experimental animal models. Metabolism of benzene is necessary for the compound’s toxic 

effects to develop. Evidence has accumulated indicating that oxidation of benzene by CYP2E1 

in the liver is the first step in initiation of benzene toxicity. Convincing evidence of the 

importance of CYP2E1 was provided by Valentine et al. (1996), who showed that genetic 

knockout mice lacking expression of the CYP2E1 protein produced much lower levels of 

benzene metabolites and failed to develop signs of genotoxicity and hematotoxicity following 

acute benzene exposure at dose levels that resulted in severe genotoxicity and cytotoxicity in 

both wild-type and B6C3F1 mice. 

The majority of benzene metabolism occurs in the liver, but the bone marrow is the target 

organ where its toxicity is expressed with the greatest sensitivity. The major hepatic metabolites 

of benzene are phenol, catechol, and hydroquinone. Catechol and hydroquinone have been 

shown to accumulate in bone marrow after benzene exposure (Rickert et al., 1979). The bone 

marrow has high peroxidase activity, which results in oxidation of the phenolic metabolites 

produced in the liver to the highly reactive 1,4-benzoquinone (Smith et al., 1989). Other target 

tissues are also characterized by high peroxidase activity (Low et al., 1995). The metabolic basis 

for the toxicity of benzene has been extensively studied (Snyder and Hedli, 1996). 

The most frequently observed toxic effect of benzene, both in humans and test animal 

models, is bone marrow depression, which leads to lymphocytopenia, leukocytopenia, 

thrombocytopenia, anemia, and aplastic anemia (Aksoy, 1991; Goldstein, 1988; Dosemeci et al., 

1996). The most sensitive effect observed in humans is the depression of ALC in peripheral 

blood (Rothman et al., 1996a). In test animal studies, the most sensitive effects observed are 

depressions of the colony-forming ability of bone marrow progenitor cells. These cells are 

responsible for producing the blood cells needed to replace the aging blood cells in the 

circulatory system. The regulation of hematopoiesis is a dynamic process in which stem and 

progenitor cells, in conjuction with bone marrow stroma, give rise to mature blood cells. The 

survival and proliferation of these cells are controlled by multiple growth factors or cytokines 

that regulate hematopoiesis. Several studies in mice and human bone marrow cultures have 

shown that benzene alters cytokine production or response to cytokines (Dempster and Snyder, 

1990; Farris et al., 1993; Cronkite et al., 1989; Irons et al., 1992; Irons and Stillman, 1993, 1996; 

Rothman et al., 1996b) 

142
Although a large number of human and experimental animal studies have been 

conducted, there are few human studies with reliable estimates of exposure to benzene and few 

long-term, repeated-dose experiments in test animals. Human studies frequently are also 

complicated by exposure to other solvents. The long-term test animal studies have used 

exposure levels that were too high to establish a reliable NOAEL, as significant adverse effects 

were observed even in the lowest dose tested in all the long-term studies examined. Thus, the 

lack of reliable NOAEL values from either human or test animal studies is an area of uncertainty 

in establishing RfD and RfC values that are protective of human health. The use of BMD 

modeling obviates the need for NOAELs; however, good-quality exposure-response data from 

the low end of the observable response range are still required to reliably estimate BMDs. 

Another area of scientific uncertainty in this assessment concerns the neurotoxic effects 

of benzene. As is the case with many other organic solvents, benzene has been shown to 

produce neurotoxic effects in experimental animals and humans after short-term exposures to 

relatively high concentrations of the compound. Benzene produces generalized symptoms such 

as dizziness, headache, and vertigo, leading to drowsiness, tremor, delirium, and loss of 

consciousness. In an occupational study, Kahn and Muzyka (1973) reported that workers 

complained of frequent headaches (usually at the end of the workday), tired easily, had 

difficulties sleeping, and complained of memory loss. Overall, there is a lack of reliable 

information on dose-related neurotoxic effects under low-dose chronic exposure conditions in 

either humans or experimental animal model systems. Li et al. (1992) reported that forelimb grip 

strength and the frequency of rapid response in Y-maze running in Kunming mice was increased 

following brief inhalation exposure to 0.78 ppm (2.5 mg/m3

) benzene, but both responses were 

decreased at concentrations of 3.13 ppm (10 mg/m3

) or higher. Several experimental 

deficiencies in this study prevent its use for calculating the risk to humans, but further 

investigation of these effects could reveal neurotoxic effects of concern to human health. 

There have been a number of developmental and reproductive studies in humans and in 

test animal model systems. Several test animal studies have shown developmental effects 

exhibiting manifestations such as reductions in numbers of live fetuses, reductions in live 

weight, and minor skeletal variants, but the benzene concentrations used also caused severe 

maternal toxicity. The studies by Keller and Snyder (1986, 1988), however, demonstrated that 

exposure to low concentrations of 5, 10, or 20 ppm (16, 32, or 64 mg/m3

) benzene in utero 

during development caused changes in colony-forming hematopoietic cells. These studies were 

considered supporting, because the LOAEL of 5 ppm is below the LOAEL of 7.6 ppm 

established for hematotoxic and immunotoxic effects observed in humans (Rothman et al., 

1996a). However, the confidence in the human data is much higher than that in the limited 

hematotoxic endpoints measured in mice by Keller and Snyder (1986, 1988). Furthermore, the 

responses did not establish consistent patterns in different ages of the progeny. Also, a limited 

number of animals were examined. These limitations make this study less useful because of the 

high degree of uncertainty that these are truly adverse effects. (See section 4.2.2.2.) 

143
6.2. DOSE RESPONSE 

6.2.1. Inhalation RfC 

Quantitative estimates of human health risk as a result of low-level chronic exposure to 

benzene via the inhalation route are based on data from human occupational inhalation exposure 

and from a subchronic inhalation study in experimental animals. Hematotoxicity and 

immunotoxicity are the critical effects observed in both humans and test animals. 

The air concentration of benzene considered to be without any appreciable risk with 

lifetime chronic human exposure (the RfC) is 3 × 10-2 mg/m3

. This value was obtained by 

applying a UF of 300 to the BMCLADJ obtained from BMR modeling of the exposure-response 

data for reductions in ALC from the Rothman et al. (1996a) occupational study, using a BMR 

level of one standard deviation change from the control mean response. This RfC is in good 

agreement with the value of 9 × 10-3 mg/m3

 based on the LOAEL in the Rothman et al. (1996a) 

study, with a UF of 1000. 

The RfC of 3 × 10-2 mg/m3

 is similarly in good agreement with the value of 5 × 10-2 

mg/m3 obtained by applying a UF of 1000 to the BMCLADJ estimated from BMD modeling of the 

exposure-response data for the hematologic effect of decreased HCT in male mice from the 

Ward et al. (1985) subchronic inhalation study, using a BMR level of one standard deviation 

change from the control mean response. Finally, both of these values are in good agreement 

with the value of 6 × 10-2 mg/m3

 based on the NOAEL in the Ward et al. (1985) study, with a UF 

of 300. 

The overall confidence in this RfC assessment is medium. The Rothman et al. (1996a) 

and Ward et al. (1985) studies were both well conducted, and various methodologies for deriving 

the RfC yielded similar results for the two studies. Furthermore, the availability of good-quality 

human data for a sensitive endpoint eliminates the uncertainty associated with basing the RfC on 

experimental animal data. However, with continuous endpoints such as the hematologic 

parameters measured in these studies, there is uncertainty about when a change in a parameter 

that has inherent variability becomes an adverse effect. Other uncertainties explicitly recognized 

in the quantitative derivation include intraspecies variability, that is, the need to accommodate 

sensitive human subgroups; the extrapolation of subchronic results to a lifetime exposure 

scenario; and database deficiencies. 

6.2.2. Oral RfD 

Quantitative estimates of human health risk as a result of low-level chronic exposure to 

benzene via the oral route are based on data from human occupational inhalation exposure and 

from a chronic gavage study in experimental animals. Once again, hematotoxic responses are 

the critical effects observed. 

The human chronic dose of ingested benzene considered to be without any appreciable 

risk (the RfD) is 4 × 10-3 mg/kg/day. This value was obtained by applying a UF of 300 to the 

oral equivalent dose extrapolated from the BMCLADJ obtained from BMD modeling of the ALC 

data from the human occupational study of Rothman et al. (1996a) using a BMR level of one 

144
standard deviation change from the control mean response, as for the RfC discussed above. This 

RfD is in good agreement with the value of 1 × 10-3 mg/kg/day based on the oral equivalent 

LOAEL from the Rothman et al. (1996a) study, with a UF of 1000. 

This RfD of 4 × 10-3 mg/kg/day is also in good agreement with the values of 7 × 10-4 

mg/kg/day and 6 × 10-3 mg/kg/day, which are 1/1000 of the BMDLADJ derived from the male rat 

ALC data in the NTP (1986) chronic gavage study and 1/3000 of the LOAEL from the NTP 

(1986) study, respectively. 

The overall confidence in this RfD assessment is medium. The Rothman et al. (1996a) 

study was well conducted, and the availability of good-quality human data for a sensitive 

endpoint eliminates the uncertainty associated with basing the RfD on experimental animal data. 

However, with continuous endpoints such as hematologic parameters, there is uncertainty about 

when a change in a parameter that has inherent variability becomes an adverse effect. Other 

uncertainties explicitly recognized in the quantitative derivation include intraspecies variability, 

to accommodate sensitive human subgroups; the extrapolation of subchronic results to a lifetime 

exposure scenario; and database deficiencies. 

A further uncertainty in the RfD ensues from the use of route-to-route extrapolation to 

estimate oral equivalent doses from inhalation exposures resulting from analysis of the Rothman 

et al. (1996a) occupational data. In experiments conducted to compare the metabolite doses to 

the target organ following oral or inhalation exposure, Sabourin et al. (1987, 1989) found that 

there was no simple relationship between the two routes of exposure. Oral doses and inhalation 

exposures that produced similar concentrations of one metabolite in the blood produced very 

different doses of another metabolite. The target specificity of benzene toxicity for the bone 

marrow progenitor cells irrespective of route of administration, however, is well documented 

both in humans and experimental animal models. Thus, route-to-route extrapolation is justified 

and introduces a lower degree of uncertainty than extrapolating from test animals to humans 

(U.S. EPA, 1999a). Use of a modifying factor of 3 to account for uncertainty in the route-toroute extrapolation was considered; however, it was deemed unnecessary. The RfD is based on 

human data for a sensitive endpoint; thus, it was felt that the composite UF of 300 provides 

sufficient protection. 

145
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