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. Environm ental Protection Agency W a shington, DCDISCLAIMER

This docum ent has been reviewed in accordance with U.S. Environm ental Protection Agency policy and approved for publication. Mention of trade nam e s or com m e rcial products does not constitute endorsem ent or recom m endation for use. Note: T his docum ent m a y undergo revisions in the future. T he m o st up-to-date version will be m a de available electronically via the IRIS Hom e Page at http://www.epa.gov/iris.

iiCONTENTS—TOXICOLOGICAL REVIEW FOR BENZ E NE (CAS No. 71-43-2)

F O R E W O R D

............................................................... v i i

AUTHORS, CONTRIBUTORS, AND REVIEW ERS

............................... v i i i

LIST OF ACRONYMS AND ABBREVIATIONS

................................... i x

E X E C U T I V E S U M M A R Y

.................................................... x i i

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. Derm al Absorption ...........................................5 3.2. DISTRIBUTION ....................................................8 3.2.1. Oral Exposure ...............................................8 3.2.2. Inhalation Exposure ..........................................8 3.2.3. Derm al Exposure ............................................9 3.3. METABOLISM .....................................................9 3.3.1. Metabolic Pathways .........................................1 0 3.3.2. Requirem e nt for CYP2E1 .....................................1 0 3.3.3. Toxicity of Benzene Metabolites ...............................1 2 3.3.3.1 Phenolic Products ...................................1 2 3.3.3.2. Trans,trans-Muconaldehyde ...........................1 3 3.3.3.3. Benzene Oxide ......................................1 4 3.3.4. Species, Route, and Rate Differences ............................1 4 3.3.5. Induction of CYP2E1 ........................................1 5 3.3.6. Mechanism of Toxicity .......................................1 6 3.3.6.1. Form ation of Covalent Adducts .........................1 6 3.3.6.2 Genotoxicity ........................................1 8 3.3.6.3 Oxidative Stress .....................................1 8 3.3.6.4. Inhibition of Cytokine Form ation .......................1 9 3.4. ELIMINATION AND EXCRETION ...................................1 9 3.4.1. Oral Exposure ..............................................1 9 3.4.2. Inhalation Exposure .........................................2 0 3.4.3. Derm al Exposure ...........................................2 1 3.4.4. Other Routes of Exposure .....................................2 1 3.5. PHYSIOLOGICALLY BASED PHARMACOKINETIC MODELS ...........2 2 3.6. TOXICOKINETICS SUMMARY ......................................2 6

4

. H A Z A R D I D E N T I F I C A T I O N ................................................2 8

iiiCONTENTS–TOXICOLOGICAL REVIEW FOR BENZ E NE (CAS No. 71-43-2) (continued)

4.1. STUDIES IN HUMANS .............................................2 8 4.1.1. Oral Exposure ..............................................2 8 4.1.2. Inhalation Exposure .........................................2 8 4.1.2.1. Hem a totoxicity ......................................2 8 4.1.2.2. Neurotoxicity .......................................4 3 4.1.2.3. Reproductive Toxicity ................................4 4 4.1.2.4. Developm ental Toxicity ..............................4 7 4.2. ACUTE AND CHRONIC STUDIES IN EXPERIMENTAL ANIMALS .......5 0 4.2.1. Hem atotoxicity .............................................5 1 4.2.1.1. Oral Exposure—Subchronic Studies .....................5 1 4.2.1.2. Oral Exposure—Chronic Studies ........................5 2 4.2.1.3. Inhalation Exposure—Subchronic Studies ................5 3 4.2.1.4. Inhalation Exposure—Chronic Studies ...................6 0 4.2.1.5. Effects on Stem Cell Populations .......................6 3 4.2.1.6. Sum m a ry of Principal Hem a totoxic Effects ...............7 7 4.2.2. Reproductive/Developm ental Effects ............................7 9 4.2.2.1. Reproductive Toxicity ................................7 9 4.2.2.2. Developm ental Toxicity ..............................8 3 4.2.2.3. Sum m a ry of Principal Reproductive/Developm ental T o x i c i t y E f f e c t s .....................................9 5 4.2.2.4. Mechanism s of Developm ental and Reproductive Toxicity ...9 6 4.2.3. Neurotoxicity ..............................................9 7 4.2.3.1. Oral Exposure ......................................9 7 4.2.3.2. Inhalation Exposure ..................................9 8 4.2.3.3. Sum m a ry of Neurotoxic Effects .......................1 0 1 4.2.4. Im m unotoxicity ............................................1 0 1 4.2.4.1. Oral Exposure .....................................1 0 1 4.2.4.2. Inhalation Exposure .................................1 0 3 4.2.4.3. Sum m a ry of Im m unotoxic Effects ......................1 0 4 4.3. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS AND MODE OF ACTION ..........................................1 0 4 4.4. SUSCEPTIBLE POPULATIONS .....................................1 0 9 4.4.1. Childhood Susceptibility .....................................1 0 9 4.4.2. Gender Differences .........................................1 1 0 4.4.3. Genetically Susceptible Populations ............................1 1 1 4.5. HAZARD IDENTIFICATION SUMMARY ............................1 1 3

5

. D O S E

- R E S P O N S E A S S E S S M E N T S .........................................1 1 6 5.1. INHALATION REFERENCE CONCENTRATION (RfC) .................1 1 7 5.1.1. Choice of Principal Study and Critical Effect ....................1 1 7 5.1.2. Benchm ark Dose Modeling ..................................1 1 9 5.1.3. RfC Derivation ............................................1 2 1 5.1.4. Com parison Analysis Based on the LOAEL .....................1 2 3 5.1.5. Com p arison Analysis Based on the W a rd et al. (1985) Experim ental

ivCONTENTS–TOXICOLOGICAL REVIEW FOR BENZ E NE (CAS No. 71-43-2) (continued)

A n i m a l S t u d y .............................................1 2 3 5.2. ORAL REFERENCE DOSE (RfD) ....................................1 2 7 5.2.1. Choice of Principal Study and Critical Effect ....................1 2 7 5.2.2. Conversion of Inhalation Exposure to Equivalent Oral Dose Rate ....1 2 7 5.2.3. RfD Derivation ............................................1 2 9 5.2.4. Com parison Analysis Based on the LOAEL .....................1 3 0 5.2.5. Com parison Analysis Based on the NTP (1986) Experim e ntal A n i m a l S t u d y .............................................1 3 0 5.3 DOSE-RESPONSE SUMMARY .....................................1 3 5

6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND D O S E

- R E S P O N S E ........................................................1 4 1 6.1. HUMAN HAZARD POTENTIAL ....................................1 4 1 6.2. DOSE-RESPONSE ................................................1 4 4 6.2.1. Inhalation RfC .............................................1 4 4 6.2.2. Oral RfD .................................................1 4 4

7

. R E F E R E N C E S ...........................................................1 4 6

vLIST OF FIGURES

F i g u r e 1

. M e t a b o l i c p a t h w a y s f o r b e n z e n e .........................................1 1 F i g u r e 2

. L i n e a r m o d e l o f A L C d a t a .............................................1 2 1 F i g u r e 3

. L i n e a r m o d e l o f H C T d a t a .............................................1 2 5

LIST OF TABLES

Table 1. Hem a totoxicity of benzene—occupational exposure ..........................3 0 Table 2. Median benzene urine m e tabolites, by exposure category, in a study of workers exposed to benzene in Shanghai, China, 1992 ..............................4 0 Table 3. Com parison of m ean peripheral blood counts with standard deviations, by exposure status, in a study of workers exposed to benzene in Shanghai, China, 1992 ......4 1 Table 4. Reproductive toxicity of inhaled benzene in hum ans ..........................4 5 Table 5. Developm ental toxicity of benzene—hum ans ...............................4 8 Table 6. Peripheral blood and hem a topoietic effects of benzene in anim als— inhalation exposure .........................................................5 4 Table 7. Reproductive toxicity of inhaled benzene in test anim als ......................8 0 T a b l e 8

. D e v e l o p m e n t a l t o x i c i t y o f i n h a l e d b e n z e n e i n t e s t a n i m a l s .....................8 4 Table 9. Joint effects of CYP2E1 activity and NQO1 genotype on benzene-induced h e m a t o t o x i c i t y i n C h i n e s e W o r k e r s ..............................................1 1 2 Table 10. Results of BMC m odeling of Rothm an et al. (1996a) data on benzene and ALC

. . 122 Table 11. BMD m odeling results of the NTP (1986) m a le m ouse and m a le rat lym phocyte counts, with untransform ed data ..............................................1 3 3 Table 12. BMD m odeling results of the NTP (1986) m a le m ouse and m a le rat lym phocyte counts, with transform ed dose data ............................................1 3 4 Table 13. Sum m a ry of uncertainty factors used for deriving the RfC and RfD ............1 3 7 Table 14. Sum m a ry of RfC and RfD estim ates using hum an and experim e ntal anim al data,

a s w e l l a s b e n c h m a r k d o s e m o d e l i n g a n d L O A E L / N O A E L a p p r o a c h e s ...............1 4 2

viFOREWORD

The National Center for Environm ental Assessm ent-W a shington Division, Office of Research and Developm ent, has prepared this docum ent on the Toxicological Review of Benzene to serve as a source docum ent for updating the noncancer health effects sum m ary on benzene in the Integrated Risk Inform ation System (IRIS).

In the developm ent of this docum ent, the published relevant scientific literature has been reviewed, key studies have been evaluated, and sum m ary/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. Environm ental Protection Agency’s standing guidance of several relevant risk assessm ent guidelines dealing with reproductive, developm ental, and neurotoxic effects. This draft has undergone intern al peer review and an expert external peerpanel review in October 1998 as well as a 90-day public com m ent period.

The em phasis of this docum ent is a discussion of the noncancer adverse health effects of benzene, including the no-observed-adverse-effect levels, the lowest-observed-adverse-effect levels, benchm ark dose analysis, uncertainty factors, and any other considerations used to develop the RfDs and RfCs for benzene.

viiAUTHORS, CONTRIBUTORS, AND REVIEWERS

The National Center for Environm ental A ssessm ent-W a shington Division (NCEA-W ) of the U.S. Environm ental Protection Agency’s (EPA’s) Office of Research and Developm ent (ORD) was responsible for the preparation of this docum ent. TN & Associates, Inc., under EPA Contract No. 68-C6-0024, W o rk Assignm ent No. 1-09.1, conducted the literature search and prepared prelim inary drafts of the docum ent.

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 Kim m el, NCEA-W Robert McGaughy, NCEA-W W illiam Pepelko, form erly of NCEA-W Jam e s Rowe, Office of Science Policy W oody Setzer, National Health and Environm ental Effects Research Laboratory-RTP

Other EPA Offices

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

Other Federal Offices

Dr. Nathaniel Rothm an, 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 Assessm ent, Cincinnati, OH Dr. John Keller, Consultant, Olney, MD Dr. Michele Medinsky, Consultant, Durham
, NC Dr. Robert Snyder, Departm e nt of Pharm acology and Toxicology, Rutgers University, Piscataway, NJ

viiiLIST OF ACRONYMS AND ABBREVIATIONS

AChE acetyl cholinesterase

ALC a bsolute lym phocyte count

AML acute m y elogenous leukem i a

ANOVA analysis of variance

ASA acetylsalicylic acid

AUC area under the curve

BFU-E burst-form ing units–erythroid

BMC b enchm a rk concentration

BMCL 95% lower confidence lim it on the BMC

BMD b enchm a rk dose

BMDL 95% lower confidence lim it on the BMD

BMR b enchm a rk response

CFU-C c olony-form ing units–culture

CFU-C/tibia colony-form ing granulopoietic stem cells

CFU-E c olony-form ing units–erythroid

CFU-HPP high proliferative potential colony-form ing units

CFU-S c olony-form ing 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 cytochrom e P450 2E1

EMG e lectrom yographical

EPA U .S. Environm ental Protection Agency

F344 Fischer 344

GD gestation day

ix CILIST OF ACRONYMS AND ABBREVIATIONS (continued)

GM-CFU-C granulocyte/m acrophage colony-form ing unit–culture

GM-CFU-G granulocyte/m acrophage colony-form ing unit

GPA glycophorin A

GSH glutathione

GTP guanosine triphosphate

HCT h em atocrit
Hgb h em oglobin

HSDB Hazardous Substance Data Bank

IC 50 Inhibition concentration 50%
i.p. intraperitoneal

IL-1 interleukin-1

IL-1 

###  interleukin-1alpha

IL-1 

###  interleukin-1beta

IRIS Integrated Risk Inform ation System

Kd kilodalton

LOAEL l owest-observed-adverse-effect level

LPS lipopolysaccharide

MA trans,trans -m uconic acid

MCH m ean corpuscular hem oglobin

MCV m ean corpuscular volum e

MDS m y elodysplastic syndrom e

MPO m y eloperoxidase

MUC m u conaldehyde

NADPH nicotinam ide adenine dinucleotide phosphate

NCEA National Center for Environm ental Assessm ent

NCEA-W National Center for Environm ental Assessm ent–W a shington Division

NK natural killer

xLIST OF ACRONYMS AND ABBREVIATIONS (continued)

NOAEL no-observed-adverse-effect level
NQO1 NAD(P)H NAD(P)H quinone oxidoreductase

NTP N ational Toxicology Program

OR odds ratio

PBPK physiologically based pharm acokinetic

PCE polychrom atic erythrocytes

PC-PFC polyclonal plaque-form ing cells

PCV p acked cell volum e

PHA phytohem a glutinin

RBC red blood cell

RfC reference concentration

RfD reference dose

SD standard deviation

SE standard error

SRBC sheep red blood cell

TW A tim e-weighted average

UF uncertainty factor

U.S. EPA U.S. Environm ental Protection Agency

W B C w hite blood cell

YAC yeast artificial chrom osom e

xiEXECUTIVE SUMMARY

The U.S. Environm ental Protection Agency’s (EPA’s, the Agency’s) Office of Mobile Sources, Office of Air and Radiation, requested that the National Center for Environm ental Assessm ent (NCEA) provide an updated characterization of hum an 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 docum ents serve as a source of inform ation regarding the carcinogenic effects of exposure to benzene.

The scope of the present report is lim ited to an assessm ent of the noncancer effects of benzene under the Agency’s standing guidance of several relevant risk assessm ent guidelines dealing with reproductive, developm ental, 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 docum ent for hazard identification and dose-response assessm ent in updating the noncancer health effects sum m ary on benzene in the Integrated Risk Inform ation System (IRIS).

Benzene, also known as benzol, is widely used as an industrial solvent, as an interm ediate in chem ical syntheses, and as a com ponent of gasoline; hence, the potential for hum an exposure is great. Inhalation exposure is the m a jor route of exposure to benzene, although oral and derm al routes are also im portant. The toxicokinetics (absorption, distribution,

m e tabolism
, and elim ination) of benzene have been studied in hum ans and experim e ntal anim al species. B enzene is readily absorbed by both test anim als and hum ans and is distributed am ong several body com p artm ents. T he parent com pound 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 m a jor route of elim ination of unm etabolized benzene in hum ans is via exhalation. Absorbed benzene is

m e tabolized to phenol and m u conic acid, followed by urinary excretion of conjugated sulfates and glucuronides. Lim ited data exist on excretion of benzene in hum ans following derm al
exposure. P hysiologically based pharm acokinetic m odels have been developed and are being im proved upon to better define interactions of benzene m e tabolism
, toxicity, and dosim etry. These interactions for hum ans exposed to low concentrations can be assessed only when the

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

m odeling fram e work. However, the current m odels are insufficiently developed and validated to allow them to predict with certainty the relationship between m e tabolism and toxicity of benzene.

Benzene exposure results in adverse noncancer health effects by all routes of adm i nistration to test anim al species. Hem a totoxicity has been consistently reported to be the m o st sensitive indicator of noncancer toxicity both in lim ited studies in hum ans and experim e ntal anim als, with bone m a rrow as the principal target organ. Chronic exposure to benzene results in progressive deterioration of hem a topoietic function. W h ether the hem a totoxic and carcinogenic effects of benzene are due to a com m on m echanism has not been established. Although leukocytopenia has been consistently shown to be a m o re sensitive indicator than anem ia of benzene toxicity in experim e ntal anim als, lym phocytopenia has been shown to be even m o re sensitive than

xiileukocytopenia. A decrease in absolute lym phocyte count (ALC) was observed to be the m o st sensitive indicator of benzene exposure in several epidem iologic studies.

Benzene has also been shown to produce neurotoxic effects in test anim als and hum ans after short-term exposure, but at relatively high concentrations; however, long-term neurotoxicity exposure studies are lacking. There is som e evidence of adverse reproductive and developm ental effects due to benzene exposure from hum an epidem iologic studies, but data are not conclusive enough to link low exposure levels to effects. No data from hum an studies were found to indicate that children are m o re sensitive than adults to benzene toxicity, nor were any significant gender differences found. The m ost frequently observed gender difference in test anim als is a greater sensitivity of m a le m i ce to benzene exposure; in rats, fem a les appear to be m o re sensitive than

m a les. However, in the National Toxicology Program (NTP) study (NTP, 1986) the m a le was the

m o st sensitive sex in the rat.

Although a large num ber of epidem iologic and experim e ntal anim al studies are available for evaluating noncancer health effects, there are few reports of hum an data with sufficiently reliable estim ates of exposure to benzene and few long-term
, repeated-dose experim e nts in test anim als. Many of the epidem iologic studies are also com p licated by exposure to other solvents, and som e of the longer-term anim al studies have em ployed exposure levels that were too high to establish reliable no-observed-adverse-effect level (NOAEL) values.

The epidem iologic occupational inhalation study by Rothm an 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 m e dian 8-hour tim e-weighted average (TW A ) concentration of 13.6 ppm (43.4 m g /m 3 ). In a subgroup of 11 workers exposed to a m e dian 8-hr TW A concentration of 7.6 ppm (24 m g /m 3 ), only ALC was still significantly reduced. Thus this concentration is a lowest-observed-adverseeffect level ( LOAEL) for benzene im m unotoxicity in hum ans. The findings from this study also indicate that white blood cells were significantly decreased and the m ean corpuscular volum e was significantly increased in the total exposed group of 44 workers occupationally exposed to a m e dian 8-hour TW A of 31 ppm (99 m g /m 3 ) in com p arison to an age- and sex-m atched control group. These dose-response effects are consistent with the well-known hem a totoxic effects of benzene in hum ans and experim e ntal anim als, as discussed in Sections 4.1.2.1 and 4.2.1 and sum m arized in Tables 1 and 6.

Reduction in ALC was the m o st sensitive endpoint in the Rothm an et al. (1996a) study, and the ALC exposure-response data were used in benchm ark dose (BMD) m odeling 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 m odels for continuous data; then, the continuous linear m odel provided the best fit. In the absence of a clear definition for an adverse effect for this continuous endpoint, a default benchm ark response of one standard deviation change from the control m ean response was selected, as suggested in EPA’s draft Benchm ark Dose Technical Guidance Docum e nt (U.S. EPA, 2000b). This default definition of a benchm ark response for continuous endpoints corresponds to an excess risk of approxim a tely 10% for the proportion of individuals below the 2 nd percentile (or above the 98 th percentile) of the control distribution for norm ally distributed effects. The BMD m odeling thus resulted in a benchm ark concentration (BMC) of 13.7 ppm (8-hour TW A) and a BMCL (95% lower confidence unit on the BMC) of 7.2 ppm (8-hour TW A).

xiiiAs 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 m g /m 3 and adjusting for continuous exposure, a BMCL ADJ of 8.2 m g /m 3 was obtained. Dividing this value by an overall uncertainty factor (UF) of 300 yields a chronic inhalation RfC of 3 × 10 -2 mg / m 3

. B ecause 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/developm ental toxicity study for benzene—com p rise the rem a inder of the 300 com posite UF.

For com p arison, an RfC was also calculated from the LOAEL of 7.6 ppm (8-hour TW A) from Rothm an et al. (1996a). After converting to m g /m 3 and adjusting for continuous exposure, a LOAEL ADJ of 8.7 m g /m 3 was obtained. Dividing this value by an overall UF of 1000 yields an RfC of 9 × 10 -3 mg / m 3

. T he UF of 1000 was based on factors of 10 to account for the use of a LOAELbecause 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 / m 3 based on the LOAEL is in good agreem ent with the result of 3 × 10 -2 mg / m 3 based on the BMCL.

Support for this chronic inhalation RfC has been provided by the experim e ntal anim al study of W a rd et al. (1985). The subchronic inhalation study of W a rd et al. (1985) was selected as a supporting study because it was the experim e ntal anim al study with the longest inhalation exposure duration and it provided good dose-response data. The exposure-response relationships for the different hem a tologic endpoints in m a le m i ce (the m o st sensitive sex/species in this study) were

m odeled using a BMD m odeling approach, and decreased hem a tocrit (i.e., volum e percentage of erythrocytes in whole blood) was chosen as the critical effect. The continuous linear m odel was selected because it provided the best fit. As above, a default level of one standard deviation change from the control m ean response was used as the benchm ark 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 / m 3 and adjusting for continuous exposure, a BMCL ADJ of 48.5 m g /m 3 was obtained. Asdiscussed 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 BMCL ADJ by the overall uncertainty factor of 1000 yields a chronic inhalation RfC of 5 × 10 -2 mg / m 3

. T his value is in good
agreem ent with the RfC of 3 × 10 -2 mg / m 3 based on BMD m odeling of the ALC data from the Rothm an et al. (1996a) hum an study.

Sim ilarly, for com p arison purposes, a chronic inhalation RfC can be derived from the NOAEL of 30 ppm observed for hem a tologic effects in the W a rd et al. (1985) study. First, the NOAEL is converted to m g /m 3 and adjusted to equivalent continuous exposure, yielding 17.1
mg / m 3

. UFs are identical to those em ployed 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 / m 3

. T his value is also in good agreem ent with the RfC derived from the Rothm an et al. (1996a) study.

xivTo derive the RfD, the sam e BMCL of 7.2 ppm (8-hour TW A) for the default benchm ark response of one standard deviation change from the control m ean response for the critical effect of reduced ALC in Rothm an 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 BMCL ADJ -equivalent oral dose rate of 1.2 m g /kg/day was obtained. Dividing the BMCL ADJ -oral of 1.2 m g /kg/day by an overall UF of 300 yields a chronic oral RfD of 4 × 10 -3 m g /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 com p rise the rem a inder of the 300 com posite UF.

For com p arison, an RfD was also calculated from the LOAEL of 7.6 ppm (8-hour TW A) from Rothm an et al. (1996a). After unit conversion, correction for continuous exposure, route-toroute extrapolation, and division by a com b ined UF of 1000, an RfD of 1 × 10 -3 m g /kg/day was derived from the LOAEL. The com b ined 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 data base deficiencies. This RfD value of 1 × 10 -3

m g /kg/day is in good agreem ent with the value of 4 × 10 -3 m g /kg/day calculated from the BMCL.

For further com p arisons, RfDs were calculated on the basis of the NTP’s chronic experim e ntal anim al study (NTP, 1986). In this study, Fischer 344 (F344) rats and B6C3F1 m i ce of both sexes were adm i nistered benzene by gavage, and blood was drawn from subgroups of anim als at various tim e points. This study identified a LOAEL of 25 m g /kg for leukopenia and lym phocytopenia in fem a le F344 rats and m a le and fem a le B6C3F1 m i ce; no NOAEL was identified. Reduction in lym phocyte count was selected as the critical effect, and the dose-response relationships at different tim e points were m odeled using a BMD m odeling approach. As above, a default benchm ark response of one standard deviation change from the control m ean was used to define the BMDs. The m odeling results suggested that the m a le rat was the m o st sensitive sex/species in this NTP gavage bioassay, and a BMDL (95% lower confidence lim it on the BMD) of 1 m g /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 com posite UF of 1000. This com posite 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 m g /kg/day, which is in reasonably good agreem ent (within an order of m a gnitude) with the RfD of 4 × 10 -3 m g /kg/day derived from the Rothm an et al. (1996a) hum an inhalation study.

An RfD was also derived from the LOAEL of 25 m g /kg observed in the NTP (1986) study. The LOAEL was adjusted to a continuous exposure level of 17.9 m g /kg/day and then divided by a UF of 3000 to derive an RfD of 6 × 10 -3 m g /kg/day. The com b ined 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 m g /kg/day is in good agreem ent with the value 4 × 10 -3 m g /kg/day derived from the Rothm an et al. (1996a) hum an study.

xv1. INTRODUCTION

The overall purpose of this docum ent 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), benchm ark dose (BMD)

m odeling, and uncertainties in derivation of an inhalation reference concentration (RfC) and an oral reference dose (RfD). This toxicological review docum ent m a y serve as the scientific basis for establishing a national air quality standard for am bient air exposure to benzene by the U.S. Environm ental Protection Agency’s (EPA’s) Office of Mobile Sources, Office of Air and Radiation, under the Clean Air Act Am endm ents of 1990, Subchapter II (42 U.S.C. 7521-7590) for hazardous air pollutants.

This docum ent presents background and justification for the noncancer health hazard and dose-response assessm ent sum m aries for benzene exposure in EPA’s Integrated Risk Inform ation System (IRIS). IRIS sum m aries include an oral RfD, an inhalation RfC, and a carcinogenicity assessm ent.

The RfD is based on the assum p tion that thresholds exist for certain toxic effects, such as cellular necrosis, but m a y not exist for other toxic effects, such as som e carcinogenic responses. It is expressed in units of m illigram s per kilogram s per day (m g/kg/day). In general, the RfD is an estim ate (with uncertainty spanning perhaps an order of m a gnitude) of a daily exposure to the hum an population (including sensitive subgroups) that is likely to be without an appreciable risk of deleterious noncancer effects during a lifetim e. The inhalation RfC is analogous to the oral RfD, but it provides a continuous inhalation exposure estim ate. The inhalation RfC considers toxic effects for the respiratory system (portal-of -entry effects) and for system s peripheral to the respiratory system (extrarespiratory or system ic effects). I t is generally expressed in units of

m illigram s per cubic m e ter (m g/m 3 ). 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.

Developm ent of these hazard identification and dose-response assessm ents for benzene has followed the general fram e work for risk assessm ent, as set forth by the National Research Council (1983). EPA guidelines that were used in the developm ent of this assessm ent 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 em ployed for this com pound were based on the Chem i cal Abstracts Registry Num b er (CASRN) and at least one com m on nam e

. A t a m i nim u m , the

1following databases were searched: RTECS, HSDB, TSCATS, CCRIS, GENETOX, EMIC, EMICBACK, DART, ETICBACK, TOXLINE, CANCERLINE, MEDLINE, and MEDLINE backfiles. Any pertinent scientific inform ation subm itted by the public to the IRIS Subm ission Desk was also considered in the developm ent of this docum ent.

2. CHEMICAL AND PHYSICAL INFORMATION RELEVANT TO ASSESSMENTS

Benzene also is known as benzol. Som e relevant physical and chem ical properties of benzene are listed below.

• C ASRN: 71-43-2

• U.S. EPA Hazardous W a ste No.: U019 (in com m e rcial product) F005 (nonspecified source, spent solvent)

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

• Synonym s: annulene, benzeen (Dutch), benzen (Polish), benzine, benzol, benzole, benzolo (Italian), bicarburet of hydrogen, coal naphtha, cyclohjexatriene, fenzen (Czech),

m i neral naphtha, m o tor benzol, NCI-C55276, phene, phenyl hydride, pyrobenzol, pyrobenzole

• R egistered trade nam e : P olystream

• E m p irical form ula: C 6 H 6

• Molecular weight: 78.11 (Budvari, 1989)

• V apor pressure: 75 m m Hg at 20°C (NFPA, 1994)

• W a ter solubility: 1750 m g /L at 25°C (Banerjee et al., 1980). Miscible with ethanol, ethyl ether, acetone, and chloroform

.

• Partition coefficients: L og K
ow : 2.13 (Hansch and Leo, 1985) Log K
oc : 1.8 to 1.9 (HSDB, 1997)

• H enry’s law constant: 5.5 × 10 -3 atm

- m 3 /m ol (Mackay and Leinonen, 1975) at 25°C

• C onversion factor: 1 ppm = 3.24 m g /m 3 at 20°C; 1 m g /m 3 = 0.31 ppm ; 1 m g /L = 313 ppm

• M elting point: 5.5 o C (HSDB, 1997)

• B oiling point: 80.1 o C at 760 m m Hg (HSDB, 1997)

• Odor: Arom atic (HSDB, 1997) Lower 0.84 ppm (HSDB, 1997) Upper 53 ppm (HSDB, 1997)

2• T aste threshold: 0.5 to 4.5 m g /L (HSDB, 1997)

• P EL (perm issible exposure lim it) TW A = 1 ppm (OSHA, 1987)

• TLV (threshold lim it value) TW A = 10 ppm (ACGIH, 1996)

3. TOXICOKINETICS RELEVANT TO ASSESSMENTS

The absorption, distribution, m e tabolism and excretion of benzene have been intensively investigated in several experim e ntal anim al species and in hum ans. Benzene is readily absorbed from oral and inhalation exposures. Derm al absorption is also rapid; however, quantitatively, derm al absorption is very low due to rapid evaporation from skin. Benzene is rapidly distributed throughout the body after exposure by all routes, and accum u lation in fatty tissues is observed. Metabolism of benzene is necessary for the expression of the characteristic hem a totoxic and carcinogenic effects of benzene. Despite extensive research, no single m e tabolite has been identified as responsible for all the toxic effects of benzene, and the weight of evidence points toward an interaction of several m e tabolites. At low doses, benzene is rapidly m e tabolized and excretion occurs prim a rily as conjugated m e tabolites; however, at higher doses m e tabolic pathways becom e saturated, and exhalation of unm etabolized benzene is observed to be the prim a ry route of excretion. Therefore, extrapolation from results observed at high doses underestim ates the potential toxic effects of low doses. The pathways of benzene m e tabolism appear to be qualitatively sim ilar across species. Quantitative differences in m e tabolism am ong anim al species are observed, however, and no good m odel for hum an m e tabolism has been established. Thus, despite the developm ent of several physiologically based pharm acokinetic (PBPK) m odels, extrapolation of results from laboratory anim al results to hum ans has proved difficult.

3.1. ABSORPTION

Benzene is readily absorbed by both test anim als and hum ans from inhalation, oral, and derm al exposures (U.S. EPA, 2000a).

3.1.1. Gastrointestinal Absorption

Although only lim ited data are available on absorption of benzene in hum ans by the oral route, accidental or intentional poisoning case studies indicate that benzene is readily absorbed (Thienes and Haley, 1972).

Nearly com p lete absorption of orally adm i nistered benzene has been dem onstrated in laboratory anim al studies. Parke and W illiam s (1953) perform ed a m a ss-balance study using radiolabeled 14 C-benzene with gavage exposures in rabbits at doses of 340–500 m g /kg. Calculation of the m a ss balance 2–3 days after exposure showed that 84–89% of the adm i nistered 14 C could be accounted for as m e tabolites, CO 2, and exhaled unchanged benzene. Thus, approxim a tely 90% of the dose was absorbed. Sabourin et al. (1987) sim ilarly found that gastrointestinal absorption of 14 C-benzene adm i nistered to rats and m i ce at a dose range of 0.5 to 150 m g /kg was greater than 97%. Although this study used corn oil as a vehicle, it is reasonable

3to assum e that oral absorption from water solutions would also be close to 100%. One could expect that the presence of food in the stom ach m a y delay absorption and an em pty stom ach m a y enhance it.

The presence of soil had only m i nor effects on absorption of orally adm i nistered 14 Cbenzene (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 ab sorption of benzene, as less than 2% of the adm i nistered dose was excreted in the feces within 48 hours. Adsorption of benzene to soil apparently increased the rate of gastrointestinal absorption, as the peak plasm a benzene concentrations were higher in both soil treatm e nts, and the tim e to peak was reduced in the sandy soil. The presence of both soils also increased the proportion of 14 C excreted in the urine, as opposed to excretion as unm etabolized benzene in expired air. Increased excretion in the urine also indicated that the proportion of the benzene subject to m e tabolism was increased. Phenol was the prim a ry m e tabolite detected in the acid-hydrolyzed urine in all treatm e nt groups, followed by hydroquinone, catechol, and benzenetriol. N o significant differences in the distribution of m e tabolites am ong the treatm e nts were found. Because the urine extracts were acid-hydrolysed, conjugation products were not determ ined. G astrointestinal absorption was apparently rapid and efficient, even in the presence of soil.

3.1.2. Respiratory Absorption

There is a significant experim e ntal database on the respiratory absorption of benzene in hum ans (Nom iyam a and Nom i yam a , 1974; Pekari et al., 1992; Srbova et al., 1950; Yu and W e isel, 1998). Srbova et al. (1950) exam ined absorption of benzene by 23 hum an subjects exposed to a range of concentrations, from 47 to 100 ppm (150 to 320 m g /m 3 ), for 2–3 hours. Absorption was greatest in the first 5 m i nutes of exposure (70–80%) but declined rapidly over the next 15 m i nutes 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. Nom i yam a and Nom i yam a (1974) determ ined both retention and uptake of benzene by three fem a le and three

m a le subjects, 18–25 years of age, exposed to 52–62 ppm (166–198 m g /m 3 ) for 4 hours. Exhaled air was sam p led every hour to determ ine respiratory excretion of benzene. Retention declined from approxim a tely 50% in the first hour and stabilized at 30% after 3 hours. Respiratory uptake averaged 47%, with excretion of 17%. No significant differences between

m a les and fem a les were reported; however, as noted, only three subjects of each sex were

exam ined.

Pekari et al. (1992) studied respiratory absorption of benzene in three m a les exposed to 1.7 and 10 ppm (5 and 39 m g /m 3 ) benzene for 4 hours each. Absorption of benzene was determ ined by m easuring 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 W e isel (1998) m easured the uptake of benzene from sidestream tobacco

sm oke by three fem a le subjects. Benzene concentrations were m easured in inhaled and exhaled air. Sm oke was generated by burning cigarettes, resulting in a variable benzene concentration ranging from 32 to 69 ppm (102 to 220 m g /m 3 ). Absorption in eight experim e nts averaged 64%, with a range of 48–73%. The exposure periods were either 30 or 120 m i nutes, but no significant decrease in absorption with the longer exposure period was observed.

4Considered together, these studies indicate that the respiratory absorption of benzene in hum ans is approxim a tely 50%. The observed decline in absorption with increasing exposure tim e is apparently due to respiratory excretion of unm etabolized benzene. Anim al studies indicate that the m e tabolism of benzene is saturated at exposure concentrations in excess of 10
ppm (32 m g /m 3 ). Thus, with the exception of the study by Pekari et al. (1992), benzene

m e tabolism m a y have been saturated in the hum an respiratory absorption studies. This would have been expected to lead to greater respiratory excretion of unm etabolized benzene. Because absorption at low dose levels is m o st relevant for derivation of an RfD, the estim ates of 48–52% retention derived by Pekari et al. (1992) are the m o st relevant values for evaluating absorption in hum ans.

A num ber of test anim al 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 m g /m 3 ]) and the equilibrium concentration of benzene in the blood of dogs. A steady-state blood level was attained within 30 m i nutes at these exposure concentrations. Exposure concentration appears to affect the retention of inhaled radioactivity, as dem onstrated by Sabourin et al. (1987) in rats and m i ce. The retention of benzene by rats and m i ce during a 6-hour exposure decreased as exposure concentration increased: 33 ± 6% to 15 ± 9% for rats, and 50 ± 1% to 10 ± 2% for m i ce, as exposure concentration increased from 26 to 2600 m g /m 3 (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 m i ce. At exposure concentrations below 350 m g /m 3 , m i ce retained approxim a tely 50% m o re radioactivity per kilogram body weight than did rats, but there was no significant difference at the highest (2500 m g /m 3 ) concentration. In general, m i ce inhaled greater am ounts of benzene per kilogram body weight due to their higher relative m i nute volum e 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 m e tabolism for benzene than either rats or m onkeys. A fter a 6-hour exposure to 7–10 ppm benzene, m i ce retained 20% of the dose, com p ared with 3–4% for rats and m onkeys. At higher exposure levels the m e tabolism becom e s saturated and a sevenfold increase in exposure concentration, between 130 and 925 ppm , results in only a threefold increase in

m e tabolism

. W h en oral doses exceed the capacity for benzene m e tabolism
, greater concentrations are then exhaled.

3.1.3. Dermal Absorption

Studies of both hum ans and experim e ntal anim als 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 experim e ntal anim als than in hum ans. The percutaneous absorption of benzene has been studied in hum ans (Franz, 1984) and in laboratory anim als (Maibach and Anjo, 1981; Franz, 1983; Susten et al., 1985). Derm al absorption is m i nim a l when

com p ared 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 am ount applied to the skin without accounting for volatilization losses, then percentage absorption figures are low and usually less

5than 1%. Although this represents realistic exposure conditions for m o st situations, absorption can be underestim ated in situations where contact with the benzene source is m a intained for a prolonged period. A substantial portion of the absorbed benzene is excreted through the lungs. Thus, experim e nts that m easure excretion only in urine and feces substantially underestim ate true absorption.

Franz (1984) studied both in vivo and in vitro derm al absorption of benzene in hum ans. A dose of 0.0026 m g/cm 2 of 14 C-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. A bsorption of the applied dose was estim ated by urinary excretion of 14 C for 36 hours. No correction was m a de for volatilization, but a correction was m a de for the fact that not all absorbed benzene is excreted in the urine. Correction was determ ined by the fraction of 14 C from a subcutaneous injection in rhesus m onkeys 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 estim ated to be 0.05% of the applied dose. Absorption using in vitro techniques with hum an skin was estim ated to be 0.1% at a sim ilar dose rate. Absorption in vitro increased linearly with dose level and also as a function of exposure tim e.

Laitinen et al. (1994) studied occupational exposure to benzene in eight car m echanics in Finland. Blood sam p les were taken 3–9 hours after exposure. The approxim a ted benzene concentration in blood corresponding to the tim e point of 16 hours after exposure showed m u ch 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 lim it of 0.2 cm 3 /m 3 to 3.7 cm 3 /m 3 , depending on whether it was in unleaded or leaded gasoline. Com parison of m easured blood concentrations to predictions based on air m easurem e nts suggested that derm al exposure could have accounted for 68% of exposure. These m e chanics had direct derm al contact with gasoline during the frequent changing of filters and fuel pum ps. This suggests far m o re derm al exposure to benzene than exposure via the inhalation route.

In a series of experim e nts conducted in a residence with a benzene-contam inated water supply, Lindstrom et al. (1993) estim ated, by m odeling, that a dose of 281 

###  g benzene would be absorbed during a 20-m inute shower, with 60% derived from derm al absorption and 40% from inhalation absorption, based on results from one individual only. The validity of this observation is highly questionable.

Several derm al absorption studies have also been conducted with experim e ntal anim als. In rhesus m onkeys, m i nipigs, and hairless m i ce, derm al 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

m a y be noted the percent absorption of the applied dose of benzene in each of these test anim als was approxim a tely twofold to threefold higher than that of hum ans.

Derm al uptake of benzene from aqueous solutions is an im portant param e ter for evaluating the risk due to exposure to benzene-contam inated ground or surface waters, for

exam ple, during showering, bathing, or swim m ing. Morgan et al. (1991) com p ared derm al

6uptake of undiluted benzene with one-third, two-thirds, and com p letely saturated aqueous solutions of benzene. A capped skin depot (3.1 cm 2 surface area) was glued to the shaved skin of

m a le Fischer 344 (F344) rats, and 2 m L of benzene or aqueous solutions were used. Blood

sam p les were collected via an im planted 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 m g /L. Derm al 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 approxim a tely 0.18 m g /L in the two-thirds saturated treatm e nt and 0.33

m g /L in the saturated treatm e nt. B enzene was essentially com p letely absorbed, with less than 1% rem a ining in solution after 24 hours. Benzene was preferentially absorbed, because the volum e of aqueous benzene solution absorbed did not differ significantly from absorption of distilled water. Benzene was rapidly and com p letely absorbed from aqueous solutions.

Tsuruta (1989) reported that derm al absorption of benzene increased linearly with dose in hairless m i ce exposed to benzene vapors (the m i ce were attached to respirators to avoid inhalation exposure). The derm al absorption rates at exposure concentrations of 200, 1000, and 3000 ppm (639, 3195, and 9584 mg/m 3 ) were 4.11, 24.2, and 75.5 m m o l/cm 2 /hour, respectively. This is equivalent to an absorption rate of 0.31, 1.89, and 5.90 μg/cm 2 /hour, respectively. The skin absorption coefficient was 0.619 cm /hour. Using the m ouse derm al absorption data and hum an occupational exposure data, Tsuruta estim ated that skin absorption of benzene by hum ans would be 3.7% that of inhalation exposure at the sam e concentration.

Perm eability constants for derm al absorption of benzene vapors were also estim ated by McDougal et al. (1990). In this study, rats were supplied air through latex m a sks and exposed to benzene vapor at a concentration of 40,000 ppm (127,787 m g /m 3 ) for 4 hours. Blood concentration was m onitored at 0.5, 1, 2, and 4 hours. A PBPK m odel was used to estim ate the
perm eability of the vapor in rat skin as well as hum an skin; the rat and hum an perm eability constants were estim ated as 0.15 and 0.08 cm /hour, respectively. On the basis of these findings, derm al exposure studies in rats probably provide a conservative estim ate of the derm al absorption of benzene by hum ans. In an in vitro experim e nt using F344 rat skin, Mattie et al. (1994) determ ined a skin:air partition coefficient of 35 for benzene at 203 ppm (649 m g /m 3 ), with an equilibration tim e of 4 hours.

Adsorption of benzene to soils resulted in a sm all decrease in derm al absorption when

com p ared with free benzene (Skowronski et al., 1988). Radiolabeled 14 C-benzene was applied to the shaved skin of m a le Sprague-Dawley rats either as free benzene or m i xed with a sandy soil (4.4% organic m a tter) or a clay soil (1.6% organic m a tter). 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 plasm a -concentration tim e curve were observed for both soils, but the effect of the clay soil was greater. No statistically significant differences in the absorption or elim ination half-lives am ong treatm e nts were observed. The greatest excretion of

14 C was observed in the urine (45–86% of applied dose), with lesser am ounts elim inated by respiratory excretion (6–13%). Excretion in the feces was 0.2% or less. These values were corrected for volatilization losses during adm i nistration of 67%, 59%, and 39% of applied radioactivity for the free benzene, sandy soil, and clay soil treatm e nts, respectively. Although these experim e nts suggested that soil m a y reduce benzene absorption, the use of occlusion of the

7exposed skin is not realistic for the derm al exposure scenarios used in risk assessm ent. Thus, the

sm all decreases in uptake have little m eaning for evaluation of risks from soil-absorbed benzene.

3.2. DISTRIBUTION

There were no studies reported regarding the distribution of benzene in hum ans after either oral or derm al exposure. However, the anim al data available provided som e useful inform ation. Both hum an and anim al data were available following inhalation exposure.

3.2.1. Oral Exposure

Low et al. (1989, 1995) followed the distribution of 14 C-benzene adm i nistered by oral gavage in fem a le Sprague-Dawley rats. This study sought to determ ine whether the distribution of benzene or benzene m e tabolites could explain the observations of tissue-specific induction of solid tum o rs in the Zym b al gland, oral and nasal cavities, and m a m m a ry glands associated with chronic oral exposure (NTP, 1986). Following oral doses of 0.15 or 1.5 m g /kg in olive oil, the concentrations of 14 C in various organs could be assigned to three groups, with the highest concentrations in liver and kidneys, interm ediate concentrations in the blood, and lowest concentrations in the Zym b al gland, nasal cavity tissue, oral cavity tissue, m a m m ary glands, and bone m a rrow. A t dose levels of 15 m g /kg or higher, disproportionate increases were found in

m a m m a ry glands and bone m a rrow. A t the 0.15 m g /kg dose level, all of the 14 C activity found in the tissues or blood after 1 hour appeared as benzene m e tabolites, indicating that first-pass

m e tabolism in the liver was very efficient. A high proportion of the m e tabolites extracted by ethyl acetate from the Zym b al gland and the nasal cavity tissue 1 hour after adm i nistration were present as unidentified m e tabolites. Sim ilarly, a high proportion of the water-soluble m e tabolites from these tissues were also unidentified peaks that did not correspond to known m e tabolites of benzene. Elim ination of benzene-derived 14 C in all organs was biphasic. Benzene did not

accum u late in the Zym b al gland; within 24 hours after adm i nistration, radiolabel derived from

14 C-benzene in the Zym b al gland constituted less than 0.0001% of the adm i nistered dose. Thus, preferential accum u lation of benzene or m e tabolites in the Zym b al gland does not account for tissue-specific tum o r induction in this organ.

3.2.2. Inhalation Exposure

Results from test anim al studies indicate that absorbed benzene is distributed throughout several body com p artm ents. T he parent com pound 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 m g /m 3 (500
ppm ) for 6 hours were blood, 1.2 m g %; bone m a rrow, 3.8 m g %; and fat, 16.4 m g % (Rickert et al., 1979). Benzene was also found in the kidney, lung, liver, brain, and spleen. Levels of the benzene m e tabolites phenol, catechol, and hydroquinone were higher in the bone m a rrow than in blood, with phenol being elim inated m o re rapidly than catechol or hydroquinone after exposure. Ghantous and Danielsson (1986) exposed pregnant m i ce to a benzene concentration of 6400
mg / m 3 (2000 ppm ) for 10 m i nutes and found benzene and its m e tabolites 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 im m e diately following exposure.

8Benzene was rapidly distributed throughout the bodies of dogs exposed via inhalation to concentrations of 800 ppm (2556 m g /m 3 ) for up to 8 hours per day for 8–22 days (Schrenk et al., 1941). Fat, bone m a rrow, and urine contained about 20-fold the concentration of benzene in blood; benzene levels in m u scles and organs were onefold to threefold that in blood, and erythrocytes contained about twice the am ount of benzene found in plasm a

. Studies in pregnant m i ce dem onstrated that after inhalation exposure, 14 C-benzene crossed the placenta. Volatile radioactivity (unm etabolized benzene) was observed in the placenta and fetuses im m e diately after and up to 1 hour after exposure (Ghantous and Danielsson, 1986). Nonvolatile m e tabolites were also detected in the fetus, but at lower levels than in m a ternal tissues. The label peaked in fetal tissues 30 m i nutes to 1 hour after inhalation, sim ilar to the peak in m a ternal tissues. No firm ly tissue-bound m e tabolites of benzene were detected in the fetal tissues in late gestation, indicating that the m ouse fetus did not have the ability to form the reactive m e tabolites.

Sato et al. (1975) com p ared elim ination kinetics of benzene in m e n and wom e n of sim ilar
ages. E xposure was for 2 hours at 25 ppm

. T he level of benzene in the blood and the end-tidal air was different for m a les and fem a les. The shape of the decay curve was significantly steeper in the m a les. The authors attributed these results to the higher fat content of fem a les.

3.2.3. Dermal Exposure

No relevant studies were found regarding distribution in hum ans following derm al exposure to benzene, and only one study in anim als was identified, providing m i nim a l inform ation on the distribution of benzene via the derm al route.

A study of m a le rats treated derm ally with 0.004 m g /cm 2 of 14 C-benzene, with and without 1 g of clay or sandy soil, revealed soil-related differences in tissue distribution following
treatm e nt. T he 14 C activity (expressed as a percentage of initial dose per gram of tissue) 48 hours after treatm e nt 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 am ount 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 m a rrow, esophagus, pancreas, lung, heart, spleen, blood, brain, thym us, thyroid, adrenal, testes, untreated skin, and carcass.

3.3. METABOLISM

Despite extensive research, the m e tabolism 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 m e tabolic product or products responsible for the noncancer and carcinogenic effects of benzene exposure have not been clearly defined. The evidence suggests that several m e tabolites, as well as interactions between these m e tabolites, m a y be necessary to explain the toxic effects of benzene. A com p lete review of the m e tabolism of benzene is beyond the scope of this toxicological review. The reader is referred to several excellent reviews for

m o re com p lete coverage (ATSDR, 1997; Snyder et al., 1993b; Snyder and Hedli, 1996; Ross,

91996; W itz et al.,1996). This section provides an overview of benzene m e tabolism
, focusing on the issues that are relevant to the interpretation of the noncancer effects due to benzene exposure.

3.3.1. Metabolic Pathw ays

The first step in benzene m e tabolism is the form ation of the epoxide, benzene oxide, catalyzed by cytochrom e P450 2E1 (CYP2E1) (Figure 1). After form ation of benzene oxide, the

m e tabolism of benzene branches into several alternative m e tabolic pathways (Jerina et al., 1968; Lovern et al., 1997). However, although Jerina et al. (1968) provided a kinetically plausible
argum ent for the existence of benzene oxide, they did not show that it was form ed from benzene. Benzene oxide rearranges nonenzym atically to form phenol, the m a jor product of initial benzene

m e tabolism

. Alternatively, benzene oxide m a y react with glutathione (GSH) to formphenylm e rcapturic acid; undergo enzym a tic conversion by epoxide hydrolase to benzene dihydrodiol with subsequent form ation of catechol; or undergo an iron-catalyzed, ring-opening reaction to form trans,trans -m uconaldehyde (MUC) with subsequent m e tabolism to trans,trans -m uconic acid (MA). Phenol is further oxidized by CYP2E1 catalysis to hydroquinone. Further oxidation of hydroquinone to p -benzoquinone is catalyzed by m y eloperoxidase (MPO) (Sm ith et al., 1989). All of the phenolic products m a y be conjugated with sulfate or glucuronic acid, and the conjugates of phenol and hydroquinone are the m a jor benzene m e tabolites excreted in urine (Sabourin et al., 1989; W e lls and Nerland, 1991).

Mathews et al. (1998) studied the m e tabolism of 14 C-benzene in m a le F344 rats over a wide range of oral (gavage) doses, 0.02, 0.1, 0.5, or 100 m g benzene/kg body weight, in m a le B6C3F1 m i ce at oral doses of 0.1 and 100 m g /kg and in m a le ham s ters at 0.02, 0.1 and 100

m g /kg. In F344 rats, at lower doses (0.02, 0.1, and 0.5 m g /kg), greater than 95% of the dose was recovered in the urine within 48 hours, and a sm all percentage (about 3%) was recovered in the breath. At higher doses, 10 and 100 m g /kg, the percentage elim inated in the breath increased to about 9 and 50%, respectively. Excretion in the feces was a m i nor route at all doses. A sim ilar pattern of disposition of the radiolabel dose was also observed in m i ce and ham s ters. M athews et al. (1998) also exam ined the profile of urinary m e tabolites form ed. Interestingly, the percentage of prephenylm e rcapturic acid and phenylm e rcapturic acid, indicators of benzene oxide production, was relatively constant across all doses for rats (~13%), m i ce (~5%), and ham s ters (~7%). However, the percentage of hydroquinone and related conjugates ranged from about 3% at the highest dose to as m u ch as 7% at the lowest doses. A higher percentage of hydroquinone m e tabolites was seen in m i ce (~30%) and in ham s ters (~30%), but it did not appear to be dose dependent. Figure 1. Metabolic pathw ays for benz ene.

Source: Ross, 2000.

3.3.2. Requirement for CYP2E1

Oxidation of benzene by the CYP2E1 isoenzym e has been dem onstrated to be required for the expression of hem a totoxicity and genotoxicity. Adm i nistration of toluene, a com p etitive inhibitor of benzene m e tabolism
, causes a decrease in benzene m e tabolite form ation and a reduction in toxicity (Andrews et al., 1977). The prim a ry oxidation of benzene by CYP2E1

10occurs in the liver, and further m e tabolism to the final toxic com pound occurs in the target tissues. S am m e tt et al. (1979) showed that partial hepatectom y of rats dim i nished both the rate of m e tabolism of benzene and its toxicity, suggesting that a m e tabolite and/or m e tabolites form ed in the liver are necessary for toxicity. Im m unoinhibition studies in rat and rabbit hepatic

m i crosom es also have im plicated CYP2E1 as the m a jor oxidative isoenzym e involved in benzene m e tabolism (Johansson and Ingelm an-Sundberg, 1988; Koop and Laethem
, 1992). Convincing evidence that CYP2E1 activity is required for expression of benzene toxicity was provided using transgenic CYP2E1 knockout m i ce that do not express hepatic CYP2E1 activity (Valentine et al., 1996). Benzene exposure at 200 ppm (640 m g /m 3 ) for 6 hours/day for 5 days resulted in severe genotoxicity and cytotoxicity in wild-type m i ce, but no toxicity was observed in CYP2E1 knockout m i ce. Thus, the requirem e nt for CYP2E1 m e tabolism has been clearly established, but the identity of the toxic m e tabolites rem a ins uncertain.

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

m a rrow and its intra- and interspecies variability in rats, rabbits, and hum ans, because it is a target organ for several chem icals, including benzene. The data dem onstrated a presence of CYP2E1 in the bone m a rrow of all species investigated, thus supporting the hypothesis of CYP2E1-dependent local m e tabolism of several chem icals that includes benzene possibly contributing to m y elotoxicity and hem a totoxicity.

113.3.3. Toxicity of Benz ene Metabolites

The products of the phenol pathway (catechol, hydroquinone, and p -benzoquinone), MA, and benzene epoxide have been proposed as benzene m e tabolites that m a y cause the toxic effects of benzene exposure.

3.3.3.1. Phenolic Products

The production of the initial benzene m e tabolites occurs prim a rily in the liver; however, the effects of benzene toxicity are prim a rily expressed as hem a totoxicity and m y elotoxicity in the bone m a rrow (Snyder and Hedli, 1996). The evidence suggests that secondary m e tabolism in bone m a rrow is required for expression of the toxicity of benzene (Schlosser and Kalf, 1989; Subrahm anyam et al., 1990, 1991). An alternative hypothesis is that m e tabolism of benzene to toxic m e tabolites can occur in the bone m a rrow itself. Metabolism of benzene to hydroquinone in the bone m a rrow of rats has been dem onstrated, but the am ount of m e tabolites produced was

sm all (Irons et al., 1980). The presence of CYP2E1 has been detected in rabbit bone m a rrow (Schnier et al., 1989); however, CYP2E1 could not be detected in the bone m a rrow of m i ce, the species dem onstrated to be m o st sensitive to benzene hem a totoxicity (Genter and Reico, 1994). Rickert et al. (1979) observed that catechol and hydroquinone concentrations persisted in the bone m a rrow 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 m a rrow and other target tissues and subsequently activated by the action of peroxidase (Sm ith et al., 1989; Rushm ore et al., 1984; Low et al., 1995). A pot ential role for conversion of hydroquinone to p

benzoquinone by the peroxidase com ponent 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 m a jor problem with the hypothesis that phenolic m e tabolites are responsible for bone

m a rrow toxicity is that adm i nistration 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 enzym e s and benzene-oxidizing enzym e s within the liver m i ght have accounted for this result. The authors observed that phenol-conjugating enzym e s were m o re concentrated in the periportal area of the liver, the first region to absorb orally adm i nistered phenol, whereas oxidizing enzym e s were m o re concentrated in the pericentral region of the liver. This could have lead to rapid excretion of orally adm i nistered phenol before it was further m e tabolized to hydroquinone by CYP2E1 in the pericentral region of the liver. A com b ination of phenol and hydroquinone is needed to cause bone m a rrow toxicity (Eastm ond et al., 1987). Intraperitoneal (i.p.) injection of either phenol or hydroquinone alone in B6C3F1 m i ce failed to cause significant bone m a rrow toxicity; however, co-adm inistration of these m e tabolites caused a reduction in bone m a rrow cellularity with a clear dose-response curve. The presence of phenol apparently stim ulated peroxidase-dependent m e tabolism of hydroquinone to p -benzoquinone. Increased covalent binding of 14 C-hydroquinone was observed in bone m a rrow when a com b ination of phenol and hydroquinone was adm i nistered to m i ce (Subrahm anyam et al., 1990, 1991).

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

m a rkedly stim ulate peroxidase activation of hydroquinone in m u rine strom a (Ganousis et al., 1992). In addition to the effects of phenol on the activity of MPO, co-adm inistration of phenol was shown to increase the concentration of phenol and hydroquinone in the blood in B6C3F1

m i ce. Legathe et al. (1994) m easured the area under the curve (AUC) blood concentration-tim e curve for phenol and hydroquinone adm i nistered alone or in com b ination at the sam e dose levels used by Eastm ond et al. (1987). Co-adm inistration increased the phenol AUC by 1.4-fold and the hydroquinone AUC by 2.6-fold in com p arison to each com pound adm i nistered alone. Legathe et al. (1994) suggested that this resulted from saturation of the enzym e s that form sulfate and glucuronide conjugates of phenolics in the liver. These results suggest that interactions of
two or m o re phenolic m e tabolites of benzene m a y be necessary to cause the observed bone

m a rrow 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 m e tabolite responsible for bone m a rrow toxicity. Evidence supporting this hypothesis has been sum m a rized by W itz et al. (1996). MUC has been shown to cause hem a totoxicity following short-term exposures. Adm i nistration of 2 m g /kg/day MUC to CD-1

m i ce for 16 days caused significant decreases in bone m a rrow cellularity, lym phocytes, red blood cell (RBC) counts, hem a tocrit (HCT) (volum e p ercentage of erythrocytes in whole blood), and hem oglobin (Hgb) and significant increases in white blood cell (W BC) count and spleen weight (W itz et al., 1985). Snyder et al. (1989) also found that adm i nistration of MUC caused bone m a rrow toxicity in m i ce and that co-adm inistration of MUC and hydroquinone resulted in a dram a tic decrease in 59 Fe 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 dem onstrated in rabbits in the early work of Parke and W illiam s (1953), and excretion of MA was also dem onstrated in m i ce, rats, cynom olgus

m onkeys, chim panzees, and hum ans (Gad-El Karim et al., 1985; Sabourin et al., 1988a, 1989, 1992). Metabolism of MUC to MA has been shown in vivo in m i ce (W itz et al., 1990a,b). Urinary excretion of MA has been used as a sensitive biom arker of benzene exposure in hum ans. At low doses, urinary MA concentration was found to be linearly correlated with tim e-weighted average (TW A ) benzene exposure concentrations (Bechtold and Henderson, 1993; Bechtold et al., 1991). Thus, the ring-opening pathway is active in several anim al species. There is no direct evidence, however, for the in vivo form ation of MUC from benzene.

Although MUC form ation has not been dem onstrated in anim als in vivo, form ation of MUC from benzene has been dem onstrated in a m ouse hepatic m i crosom al system (Latriano et al., 1986; Zhang et al., 1995a). MUC m a y be too reactive to be isolated from in vivo system s. Using isolated rat livers perfused with 0.7 m M benzene solutions through the portal vein, Grotz et al. (1994) dem onstrated that the com p lete pathway for form ation of MA from benzene was active in the liver, but MUC was not detected in the perfusate collected via the hepatic vein. W h en MUC was added to the perfusion solution, it was rapidly and efficiently m e tabolized to MA, and only traces of MUC were detected in the perfusate collected after a single pass through the isolated rat livers. Thus, it seem s 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-

13hexadienoic acid (COOH-M-OH) and MA were the m a jor MUC m e tabolites 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 m i nor peaks (Grotz et al., 1994). The CHO-M-OH has been dem onstrated to react with GSH, to be cytotoxic to isolated rat hepatocytes, and to be hem a totoxic in m i ce (Goon et al., 1993; Zhang et al., 1995b). Thus, CHO-M-OH m a y be the ring-opened m e tabolite that causes the hem a toxic effects of adm i nistered MUC, and it m a y 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 m a rrow. H owever, Lovern et al. (1997) recently showed that benzene oxide constituted 7% of the benzene m e tabolites after 18 m i nutes’ incubation with liver m i crosom es. Lindstrom et al. (1997) dem onstrated the presence of benzene oxide in the blood and estim ated its half-life to be about 8 m i nutes. Using a PBPK m odel, 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 form ation. Further research is needed to establish the role of benzene oxide in causing bone m a rrow toxicity.

3.3.4. Species, Route, and Rate Differences

Differences in the rates of benzene m e tabolism and the m e tabolites form ed correlate with observed differences in sensitivity to benzene toxicity. Mice are m o re sensitive than rats to the toxic effects of benzene (Huff et al., 1989; Snyder et al., 1978; W a rd et al., 1985). Sabourin et al. (1987) found that m e tabolism of benzene, determ ined by excretion of water-soluble

m e tabolites in the urine, becam e saturated at lower doses in B6C3F1 m i ce than in F344/N rats. The am ount of m e tabolites per kilogram body weight was sim ilar in the two species at gavage doses up to 50 m g /kg, but above this level, total m e tabolites in rats continued to increase but total m e tabolites in m i ce did not increase further. Following inhalation exposures, total

m e tabolites excreted were higher for m i ce than for rats at all benzene concentrations due to a higher am ount inhaled by m i ce. Total m e tabolite excretion was exponentially related to benzene concentration, but the concentration needed to reach half of the m a xim a l m e tabolite form ation was lower in m i ce (220 m g /m 3 [69 ppm ]) than in rats (260 m g /m 3 [81.5 ppm ]). By both oral and inhalation routes, the m e tabolism of benzene was saturated at lower concentrations in m i ce than in rats. Thus, differences in total m e tabolite form ation between rats and m i ce did not explain the greater sensitivity of m i ce to benzene.

Sabourin et al. (1988a) com p ared the m e tabolites of benzene in the blood, liver, lung, and bone m a rrow of m a le B6C3F1 m i ce and F344/N rats following 6 hours of inhalation exposure to benzene at 50 ppm (160 m g /m 3 ). Hydroquinone glucuronide, hydroquinone, and MA were present in m u ch higher concentrations in the m ouse than in the rat tissues. Thus, m e tabolism in

m i ce leads to m u ch greater exposures to potentially toxic benzene m e tabolites, which m a y explain the greater sensitivity of m i ce to benzene toxicity.

14Exposure concentration and routes also affect the distribution of m e tabolites form ed in

m i ce and rats (Sabourin et al., 1989). Male B6C3F1 m i ce and F344/N rats were given gavage doses of 1, 10, or 200 m g /kg benzene or they were exposed to 5, 50, or 600 ppm (16, 160, or 1916 m g /m 3 ) for 6 hours. W a ter-soluble m e tabolites were determ ined in blood, urine, liver, lung, and bone m a rrow. A shift in the distribution from potentially toxic m e tabolites, hydroquinone glucuronide, and MA to the detoxification m e tabolites, phenylglucuronide and prephenylm e rcapturic acid, was observed from low concentrations to high concentrations following both oral and inhalation exposures in m i ce. A sim ilar shift was observed in rats, except that hydroquinone glucuronide was a m i nor m e tabolite in rats at all concentrations. Thus, extrapolation from toxicological studies conducted at high exposure concentrations m a y significantly underestim ate risks at lower exposure concentrations. There was also no sim p le correlation between the dose of m e tabolites to tissues resulting from oral and inhalation
exposures. This m a y be due to saturation of benzene m e tabolism following bolus doses of benzene adm i nistered by gavage.

Urinary excretion of benzene m e tabolites has also been exam ined in cynom olgus

m onkeys and chim panzees (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 m g /kg in
cynom olgus m onkeys. P henyl sulfate was the prim a ry m e tabolite excreted, and the proportion excreted as hydroquinone conjugates and MA decreased as the dose increased, as was observed previously in m i ce 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 m u ch lower in the m onkey than in either m i ce or rats. Chim p anzees were exposed at only one low dose rate of 1 m g /kg by intravenous injection. Less than 15% of the adm i nistered dose was recovered as hydroquinone conjugates and MA, and 79% was recovered as phenyl conjugates. On the basis of the urinary m e tabolite profiles, the m ouse appears to m e tabolize the largest fraction of benzene via pathways leading to hydroquinone conjugates and MA (67% at low doses) followed by m onkeys (31%), rats (17%), and chim panzees (14%). Thus, there are apparently large quantitative differences am ong species in the m e tabolism of benzene to potentially toxic m e tabolites. Because few data are available on the proportion of benzene

m e tabolized to potentially toxic m e tabolites in hum ans, considerable uncertainty exists in determ ining which anim al m odel best represents hum an m e tabolism

. Henderson (1996) reviewed the species differences in benzene m e tabolism

. The pathways of benzene m e tabolism appear to be sim ilar in all species studied; however, there are quantitative differences in the fraction of benzene m e tabolized by different pathways. Monkeys and m i ce m e tabolize m o re of the benzene dose to hydroquinone m e tabolites than do rats or chim panzees, especially at low doses. Mice appear to have a greater overall capacity to

m e tabolize benzene than do rats and prim a tes. This finding m a y explain why m i ce are m o re sensitive than rats to benzene. In all species, a greater proportion of benzene is converted to hydroquinone and ring-open m e tabolites 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 m e tabolite form ation. Pretreatm e nt of m i ce, rats, and rabbits subcutaneously with benzene increased benzene m e tabolism in vitro without increasing total cytochrom e P450

15concentrations (Arinc et al., 1991; Gonasun et al., 1973; Saito et al., 1973). In contrast, Sabourin et al. (1990) found no significant effect on the m e tabolism of benzene when F344 rats and B6C3F1 m i ce were pretreated by inhalation exposure to 600 ppm (316 m g /m 3 ) of benzene. The rate of benzene m e tabolism can be altered by pretreatm e nt with various com pounds. C YP2E1 also m e tabolizes 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 hum an hem a topoietic stem cells (Henschler and Glatt, 1995). Therefore, exposure to chem icals that stim ulate the activity of this enzym e system prior to exposure to benzene could increase the rate of benzene m e tabolism

. Daiker et al. (1996) found that repeated oral benzene exposure of fem a le B6C3F1 m i ce for 3 weeks at 50 m g /kg/day decreased CYP2E1 activity by 34% and activated the detoxification
enzym e GSH transferase by 30% without affecting aldehyde dehydrogenase, another detoxifying
enzym e

. T he authors suggested that these changes in enzym e activity m a y 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 (Nakajim a et al., 1985). The results of further studies showed that ethanol treatm e nt increased the production of hydroxylated benzene m e tabolites, phenol, and hydroquinone, suggesting induction of benzene m e tabolism (Nakajim a et al., 1987). The possibility therefore exists of a synergism between alcohol and benzene.

3.3.6. Mechanism of Toxicity

Just as several benzene m e tabolites have been im plicated in benzene toxicity, several different m echanism s also m a y contribute to the overall toxic effects of benzene, as discussed below.

3.3.6.1. Formation of Covalent Adducts

Benzene m e tabolites form covalent adducts with both cell proteins and DNA; however, the role of adduct form ation in toxicity is unclear. Treatm e nts that reduce benzene toxicity also reduce adduct form ation. Partial hepatectom y of rats resulted in reduced hem a totoxicity and also in reduced covalent binding in bone m a rrow (Sam m e tt et al., 1979). Levels of adduct form ation in the hem a topoietic tissues of m i ce strains have been shown to correlate with the relative sensitivity of the strains (Longacre et al., 1981).

Covalent binding of benzene m e tabolites to DNA was first dem onstrated by Lutz and Schlatter (1977). Mazzullo et al. (1989) found that form ation of DNA adducts of benzene

m e tabolites was linear at low benzene concentrations, but saturation of adduct form ation occurred at high benzene concentrations. Hedli et al. (1996) investigated DNA adduct form ation from the benzene m e tabolites hydroquinone and 1,2,4-benzenetriol in com b ination with investigations of the effects of these two m e tabolites on cell differentiation in a hem a topoiesis

m odel system

. H ydroquinone form ed DNA adducts in hum an prom yelocytic leukem i a cells, but 1,2,4-benzenetriol did not. Both m e tabolites, however, inhibited retinoic acid-induced

m a turation of hum an prom yelocytic leukem i a cells to granulocytes. Thus, DNA adduct form ation m a y be im portant in hydroquinone but not in 1,2,4-benzenetriol toxicity.

16Creek et al. (1997) showed that DNA and protein adduct form ation following i.p. adm i nistration of 14 C-benzene was linear over a dose range spanning eight orders of m a gnitude in B6C3F1 m i ce. At doses greater than 16 m g /kg body weight, however, adduct form ation was nonlinear. T his corresponds very closely to the m e tabolic saturation level of 15 m g /kg body weight observed by Sabourin et al. (1987). Benzene was adm i nistered to m a le B6C3F1 m i ce over a dose range of 700 pg/kg to 500 m g /kg body weight. Liver DNA adduct levels peaked at 0.5 hours after exposure, but bone m a rrow DNA adduct levels peaked between 12 and 24 hours. These results indicate that adduct form ation is linear in the range of benzene concentrations where benzene toxicity is first detected and becom e s saturated at higher doses where m o st toxicity experim e nts have been conducted.

Although the m echanism by which adduct form ation results in bone m a rrow toxicity is not well established, a few experim e nts have suggested that binding of hydroquinone or p

benzoquinone 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 polym erization during m i crotubule form ation and, therefore, interferes with spindle form ation during m itosis. Several m e tabolites of benzene have also been shown to inhibit the activity to hum an topoisom e rase II (Frantz et al., 1996). Topoisom e rase II enzym e s relieve torsional strain on DNA during replication and transcription and are also believed to function during

recom b ination and chrom osom e condensation. Both p -benzoquinone and MUC directly inhibited topoisom e rase at concentrations as low as 10 

###  M. W ith the addition of horseradish peroxidase and peroxide to the assay m i xture, the benzene m e tabolites phenol, 4,4' -biphenol, 2,2' -biphenol, hydroquinone, catechol, and 1,2,4-benzenetriol all inhibited topoisom e rase II. Topoisom e rase inhibitors tend to be strong clastogens in m a m m alian cells. Thus, inhibition of topoisom e rase II could lead to the clastogenic effects observed following benzene exposure. Addition of GSH to the assay m i xture also protects against topoisom e rase II inhibition, thus suggesting that interaction of benzene m e tabolites with an essential sufhydryl residue occurs. Hutt and Kalf (1996) also reported that p -benzoquinone inhibits the activity of topoisom e rase II.

Protein adduct form ation by benzene m e tabolites differs greatly between m i ce and rats. McDonald et al. (1994) used uniform 14 C-benzene to determ ine total protein binding and C 6 - 13 C-benzene to determ ine form ation 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 estim ated for blood Hgb and bone m a rrow total protein. Form ation 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 approxim a tely 25% of the cysteine binding in m ouse Hgb or bone m a rrow proteins of either species. Thus, other benzene m e tabolites m u st also bind to cysteine. Adducts of benzene oxide were highest in rat Hgb but accounted for only a sm all proportion of adduct form ation in m ouse Hgb. Benzene oxide adducts accounted for less than 3% of total adduct form ation in the bone m a rrow of either m i ce or rats, and benzoquinone adduct form ation accounted for a m u ch greater proportion of total protein binding in bone m a rrow. However, 1,2-benzoquinone adducts were m o re prevalent in rat bone m a rrow proteins and 1,4- benzoquinone adducts were m o re prevalent in m ouse bone m a rrow proteins. This suggests that there are significant differences in benzene m e tabolism between m i ce and rats. A background level of both 1,2- and 1,4-benzoquinone protein adducts that greatly exceeded the level of

17labeled adduct form ation was also observed in bone m a rrow proteins of control m i ce and rats. This result suggests that low doses of benzene m a y not cause toxicity through protein adduct form ation.

3.3.6.2. Genotoxicity

The association between benzene exposure and the appearance of structural and

num erical chrom osom al aberrations in hum an lym phocytes suggests that benzene m a y be considered as a hum an clastogen. In in vivo anim al studies, benzene induced cytogenetic effects, including chrom osom e and chrom atid aberrations, sister chrom atid exchanges, and

m i cronuclei (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; W a rd et al., 1992). Several lines of evidence also indicate that benzene is genotoxic in hum ans under occupational exposure conditions (Ding et al., 1983; Sasiadek et al., 1989; Yardley-Jones et al., 1990; Major et al., 1992; Eastm ond, 1993; Tom p a et al., 1994). However, these studies lacked good exposure

m onitoring data, involved m u ltiple chem ical exposures, and were often poorly designed, with inappropriate control groups.

Benzene has been shown to produce DNA breaks in Chinese ham s ter ovary cells independent of m e tabolic activators; however, in vitro assays indicate that genotoxicity of benzene is prim a rily due to its m e tabolites (Zhang et al., 1993; Eastm ond et al., 1994). Benzene is known to affect cell cycle progression, RNA and DNA synthesis, and DNA binding (Forni and Moreo, 1967). Chen and Eastm ond (1995) showed that benzene m e tabolites can adversely affect hum an topoisom e rases; however, som e DNA repairs m a y occur in hum an cells (Chenna et al., 1995).

Even with these studies, no data exist on the quantitative relationship between m easured benzene exposures and clastogenic effects.

Benzene itself is not m u tagenic in short-term assays in either bacterial or anim al system s (Dean, 1985). Mutagenicity of benzene m e tabolites, 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 ham s ter V79 cells. MUC is an active

m

u tagen in V79 cells, but it is only weakly active in Salm onella (W itz et al., 1990a).

3.3.6.3. Oxidative Stress

Benzene m a y also produce oxidative stress in the target tissues. The benzene m e tabolite
p -benzoquinone is highly reactive and can deplete cellular levels of GSH (Brunm ark and Cadenas, 1988). Benzene m e tabolites can also be involved in redox cycling, resulting in the production of reactive oxygen species that can also react with m acrom olecular com ponents (Rao and Snyder, 1995).

183.3.6.4. Inhibition of Cytokine Formation

The strom al m i croenvironm ent of the bone m a rrow that norm ally m odulates stem cell proliferation and differentiation is a potential target for the hem a totoxicity of benzene (Cox, 1991; Snyder et al., 1989; Kalf, 1987). The interaction of the strom a with the stem cells is necessary for hem a topoiesis. F urtherm ore, the strom al m acrophage produces interleukin-1 (IL1), a cytokine also essential for hem a topoiesis. P atients with aplastic anem ia usually exhibit

m onocyte dysfunction and decreased IL-1 production (Renz and Kalf, 1991).

Renz and Kalf (1991) dem onstrated the disruption of IL-1 production by the strom al

m acrophages of m i ce exposed to benzene. The m i ce were injected i.p. with 600 or 800

m g /kg/day benzene for 2 days. The strom al m acrophages rem oved from these m i ce 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 m a ture cytokine. Hydroquinone added in vitro also inhibited the conversion of the pre-IL-1 

###  to the m a ture cytokine in m ouse

m acrophages. However, adm i nistration of recom b inant m ouse IL-1 

###  to m i ce before a bonem a rrow-suppressing dose of benzene am eliorated the bone m a rrow depression, probably by circum venting the inability of the strom al m acrophage in benzene-treated anim als to process preIL-1 

###  to the m a ture cytokine. Thus, Renz and Kalf (1991) suggested that benzene-induced depression of bone m a rrow cellularity m a y result from the failure of the strom al m acrophages to process pre-IL-1 

###  to m a ture IL-1 

###  , which activates the strom al fibroblast production of the colony-stim ulating factor required for the differentiation of stem cells.

Niculescu et al. (1995, 1996) dem onstrated that p -benzoquinone, the oxidation product of hydroquinone in the cell, causes a concentration-dependent inhibition of highly purified hum an platelet calpain with an IC 50 of 3 

###  M. Calpain is a calcium -activated, cysteine-dependent protease that catalyzes the processing of pre-IL-1 

###  to the m a ture cytokine in vivo. The investigators also dem onstrated 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 inform ation base on the clearance of benzene from the body after oral exposure to the com pound is lim ited to reports of studies in experim e ntal anim als. For exam ple, in one of a series of toxicological reports on the m e tabolism and excretion of radiolabeled xenobiotics, Parke and W illiam s (1953) recovered a substantial proportion of the adm i nistered dose on the breath as unm etabolized product (43%). This am ount com p ared with 33% urinary excretion in which the label partitioned into such com ponents as conjugated phenols (23% of the dose), hydroquinone (4.8%), and catechol (2.2%), am ong others. Residual am ounts of radioactivity

rem a ined deposited in the tissues or were excreted in the feces.

In a study using m a le C57BL/6N m i ce given single oral doses of 14 C-benzene (10 or 200

m g /kg), McMahon and Birnbaum (1991) reported the effects of age on benzene m e tabolism

. Radioactivity was m onitored in urine, feces, and breath. Consistent with previous reports, the

19following urinary m e tabolites were detected: hydroquinone glucuronide, MA, phenylglucuronide, phenyl sulfate, catechol glucuronide, hydroquinone sulfate, and prephenylm e rcapturic acid. At various tim e points up to 48 hours after dosing with 10 m g /kg, a significant decrease in the urinary excretion of benzene-derived 14 C was observed for 18- versus 3-m onth-old m i ce. W h en expressed as the percent of total benzene adm i nistered, the relative

am ounts of som e urinary m e tabolites varied between the 10 and 200 m g /kg dose groups, thereby indicating dose-related quantitative changes in the urinary excretion of benzene m e tabolites. At the low dose (10 m g /kg), urinary excretion was the m a jor route of elim ination. Hydroquinone glucuronide, phenyl sulfate, and MA were the m a jor m e tabolites at this dose, accounting for 40%, 28%, and 15% of the dose, respectively. At 200 m g /kg, urinary excretion decreased to account for 42–47% of the adm i nistered dose, whereas respiratory excretion of volatile

com ponents increased to 46–56% of the adm i nistered dose. Fecal elim ination was m i nor 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 m i ce, F344 rats, and Sprague-Dawley rats via gavage to doses ranging from 0.5 to 300 m g /kg 14 C-benzene. At the lower doses (< 15 m g /kg), near quantitative am ounts of label were recovered in various urinary

m e tabolites, whereas at the higher doses, unchanged 14 C-benzene began to appear in expired air in increasing proportions. For exam ple, at doses of 150 m g /kg and above, greater than 50% of the load was cleared on the breath as unchanged starting com pound. These data point to the saturation of the m e tabolic processes with increasing dose, although even at lower doses a degree of fluctuation in the m e tabolic pattern was im plied by dose-dependent changes in the profile of m e tabolic products detected in the urine.

3.4.2. Inhalation Exposure

In a key report of studies that addressed the m e tabolic fate of inhaled benzene issue in hum ans, Nom i yam a and Nom i yam a (1974) showed that at least a proportion of the absorbed

com pound can be excreted in the urine as sulfate- or glucuronide-conjugated phenols or MAs. At the benzene concentrations em ployed (52–62 ppm [166–198 m g /m 3 ]), respiratory uptake was considered to be in the region of 47% and respiratory excretion for the 4-hour exposure period was approxim a tely 17%. These values were in reasonably good agreem ent 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 elim ination route or another as being of prim a ry im portance when hum an beings are exposed to benzene via inhalation. For exam ple, Sherwood (1988), using a com p lex experim e ntal protocol, em ployed a single hum an subject who was alternately exposed to either 6.4 ppm (20 m g /m 3 ) benzene for 8 hours or 99 ppm (316 m g /m 3 ) for 1 hour to m onitor the kinetics of benzene elim ination 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

20was biphasic, with an initial rapid excretion phase followed by a slower excretion phase. The im portance of urinary excretion was also em phasized 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 m g /m 3 ]).

Occupational exposure and cross-sectional studies have also pointed to the appearance of benzene m e tabolites in urine as a consequence of exposure to such agents as sidestream tobacco

sm oke (Bartczak et al., 1994) and gasoline vapors (Lagorio et al., 1994) or when the urine content of sm okers versus nonsm okers was com p ared (Kok and Ong, 1994; Melikian et al., 1994). In general, one cigarette delivers an average of 55 μg of benzene, and an average sm oker takes in about 1.2–1.8 m g of benzene per day from sm oking. Popp et al. (1994) detected

m

u conic acid and S-phenyl-N-acetyl cysteine levels in the urine of car m echanics at levels that correlated with levels of the com pound in the bloodstream and the breathing zone. These observations support the earlier suggestion by Ghittori et al. (1993) that benzene and its

m e tabolites in the urine m a y be an im portant biom arker of occupational exposure to the

com pound.

An accum u lation of experim e ntal data in laboratory anim als has shown an essentially sim ilar pattern of benzene elim ination and excretion as in hum an beings. In broad term s, this is characterized by the release of unchanged com pound on the breath and the appearance of

m e tabolites in the urine. For exam ple, at high incidental concentrations of benzene (500 ppm [1597 m g /m 3 ] for 6 hours), an initial rapid phase of elim ination was followed by a slower phase of m u ch longer duration (up to 13 hours or m o re) (Rickert et al., 1979).

Perhaps the m o st rigorous experim e nts that have exam ined 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 experim e ntal protocol featured a 6-hour, nose-only exposure to concentrations of 14 C-benzene ranging from 10 to 1000 ppm (32 to 3195 m g /m 3 ). 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 m g /m 3 ]), both rats and m i ce exhaled considerable proportions of the unchanged com pound, am ounting to 48% of the load in rats and 14% in m i ce. In a sim ilar experim e nt, Dow Chem i cal Co. (1992a) reported that Duroc-Jersey pigs exposed to 0, 20, 100, or 500 ppm (0, 64, 319, or 1597 m g /m 3 ) 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 lim ited data on excretion of benzene after derm al exposure in hum ans were found. Franz (1984) reported that four volunteers exposed to 14 C-benzene (0.0024 m g /cm 2 ) on the skin excreted trace am ounts of label in the urine over a 36-hour period, suggesting the ability of the

com pound to penetrate the derm al barrier. Sim ilar results were obtained in m onkeys and

m i nipigs, with excretion of the com pound in m onkeys approaching 0.065% (range 0.033–0.135%) of the applied dose, com p ared with 0.042% (range 0.030–0.054%) in m i nipigs (Franz, 1984).

3.4.4. Other Routes of Exposure

21This section discusses very lim ited data available from three studies in which benzene was either given intraperitoneally or subcutaneously. In these studies, the exposure was lim ited to either a single injection or daily injections for up to 3 days.

Experim e nts by Cornish and Ryan (1965) used the i.p. route of adm i nistration to exam ine the influence of the fed or fasted state on the rates of production of urinary m e tabolites such as conjugated or unconjugated phenols, com ponents that appeared to be increased m a rkedly in the fasted state versus fed. In line with their findings of the dose-response effects of benzene when adm i nistered via gavage or inhalation, Sabour in et al. (1987) found that the proportion of radioactivity (m ainly as phenyl sulfate) excreted in the urine decreased with increasing dose when m onkeys were injected intraperitoneally with concentrations of 14 C-benzene ranging from 5 to 500 m g /kg, again suggesting that benzene m e tabolism can be saturated at the higher dose levels. F inally, Longacre et al. (1981) exposed C57BL/6N and DBA/2 m i ce to benzene subcutaneously and m onitored the appearance of conjugated phenol, catechol, and hydroquinone in the urine, although with varying proportions when one strain was com p ared with another. Am ounts of phenyl m e rcapturic acid were sim ilar, however.

3.5. PHYSIOLOGICALLY BASED PHARMACOKINETIC MODELS

PBPK m odels can be used to address the uncertainty of extrapolating from data of an experim e ntal study to a hypothetical adm i nistered dose for the sam e response in other anim als or hum ans. The approach seeks to predict the dose-response characteristics of a chem ical’s potential toxicity through an understanding of the underlying m echanism s by which a substance is absorbed, transported to, and m e tabolized in the prim a ry target organs, on the assum p tion that the biochem i cal processes involved in the toxicological response will be the sam e for both hum ans and the anim al m odel(s) in which the harm ful effects of the com pound had been dem onstrated. N ecessary inputs to the m odel include a physiologically realistic discrim i nation of a chem ical’s m ovem e nt between m e tabolically and functionally linked com p artm ents of the body and the inform ation necessary to describe this m ovem e nt m a them atically in term s of (1) the kinetics of m e tabolism
, (2) rates of m ovem e nt to and from different target organ groupings, and (3) the partitioning of a com pound between physiological m e dia. Thus, by m odeling a target organ-specific internal dose surrogate (parent com pound or m e tabolite, as applicable) from the NOAEL or LOAEL of an experim ental study, an equivalent allom etrically scaled internal dose surrogate in hum an beings can be used to back-extrapolate to a hypothetical “effective” dose for the sam e toxicological response in hum an 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 experim e ntal data. Lim itations reside in the accuracy/validity of the estim ates that constitute the inputs to the m odel and in the oversim plification that is inherent in seeking to define the m ovem e nt and m e tabolism of a xenobiotic or its m e tabolites in term s of a sm all num ber of functionally defined subcom partm e nts.

Several attem p ts have been m a de to develop PBPK m odels to define the understanding of interactions between benzene m e tabolism and toxicity. The first m odel for benzene was developed by Sato (Sato and Nakajim a, 1979; Sato, 1988), who exposed three m e n to 25 and 100

22ppm (80 and 319 m g /m 3 ) benzene vapor for 2 hours and then observed a triexponential decay of benzene from their blood. The investigators constructed a three-com p artm ent m odel consisting of richly perfused tissues, poorly perfused tissues, and fat, which acted as a m a jor sink for benzene. Subsequently, PBPK m odels have been developed to take into account differences in benzene m e tabolism between species and individuals using both experim e ntal data and sim u lations (Medinsky et al., 1989a; Travis et al., 1990a, b; Bois et al., 1991b; Cox, 1991, 1996).

A m odel (Medinsky et al., 1989a, b, c) was developed to describe and predict the fate of benzene and to determ ine whether differences in the m e tabolic rates between rats and m i ce could explain the differences in toxicity between these species. The m odel indicated that for inhalation concentrations up to 1000 ppm (3195 m g /m 3 ), m i ce m etabolize at least two to three tim es as

m

u ch benzene as rats. However, following oral exposure, rats m e tabolized m o re benzene on a body-weight basis than did m i ce at doses greater than 50 m g /kg. Patterns of m e tabolites also differed between rats and m i ce. Mice produced prim a rily hydroquinone glucuronide and MA

m e tabolites linked to toxic effects; on the other hand, rats produced prim a rily phenyl sulfate, a detoxification product. These sim u lated results agree with experim e ntal data and provide a fram e work for understanding the greater sensitivity of the m ouse to benzene toxicity.

The Medinsky m odel was based on an earlier PBPK m odel developed by Ram sey and Anderson (1984). The tissue com p artm ents initially included in the m odel were (1) the liver,
presum e d to be the only organ where benzene m e tabolism takes place; (2) a group of poorly perfused tissues, including m u scle and skin; (3) a group of richly perfused tissues, including bone m a rrow, kidney, and intestines; and (4) a fat com p artm ent. All m e tabolism of benzene, consisting of initial m e tabolism to benzene oxide that is then further m e tabolized by one of four pathways, was m odeled by Michaelis-Menten kinetic param e ters. M etabolic rate constants were determ ined by fitting the results of m odel sim u lations to experim e ntal data obtained by exposing

m i ce 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 m odel did not sim u late the data of Rickert et al. (1979) very well, thereby indicating a need for the m odel to be further refined.

Since the original Medinsky m odel was published, additional com p artm ents have been added to reflect an advancing understanding of benzene m e tabolism
, and specific biochem i cal and toxicokinetic param e ters have been refi ned 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) m easured a 13-fold variability in CYP2E1 activity in hum an hepatic

m i crosom es and com p ared this to the activity in m ouse and rat liver m i crosom es. T he m odel predicted the dependence of benzene m e tabolism on the m easured CYP2E1 activity, and the proportion of hydroquinone (the suspected toxic m e tabolite) produced in vitro was correlated with the level of CYP2E1 activity. Seaton et al. (1995) m easured the initial rates of the two

m a jor conjugation reactions, phenol sulfonation and hydroquinone glucuronidation, in the hepatic m i crosom e preparations of hum ans, rats, and m i ce. This inform ation was used in a physiological com p artm ent m odel to predict steady-state concentrations of phenol and hydroquinone in blood.

Am ong hum ans, 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 m i ce than in rats (2.28 vs. 0.83 nM), and

23predicted hydroquinone concentrations were also higher in m i ce than in rats (42.44 vs. 17.99 nM). The predicted concentrations for m i ce were higher than the range for hum ans, but the rat values were within the predicted concentrations for hum ans. On this basis, the authors suggested that the rat m a y be a good m odel for hum ans with respect to tissue dosim etry for these benzene

m e tabolites. The authors also suggested that the m ouse m i ght be m o re sensitive than the hum an and that in vitro m e tabolism data m u st always be placed within the context of the whole anim al’s physiology.

The Medinsky PBPK m odel has served to organize the available inform ation into a coherent m odel that has helped to refine the specific experim e ntal approaches used to fill the gaps in the understanding of the m echanism of benzene toxicity. Although current PBPK

m odels m a y provide insights about putative toxic m e tabolites and potential biochem i cal

m echanism s , 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 m odel to describe the pharm acokinetics of benzene in rats, m i ce, and hum ans. The m odel contains five com p artm ents, consisting of liver, fat, bone m a rrow, m u scle, and organs such as brain, heart, kidney, and viscera. The different

com p artm ents are connected by the arterial and venous blood pathways. Metabolism of benzene is assum e d to follow Michaelis-Menten kinetics in all species and is assum e d to occur prim a rily in the liver and to a lesser extent in the bone m a rrow. M odel sim u lations were com p ared with experim e ntal data from Sabourin et al. (1987, 1988a), Andrews et al. (1977, 1979), Nom i yam a and Nom i yam a (1974), Snyder et al. (1981), Sato et al. (1975), and Rickert et al. (1979). The Travis m odel successfully sim u lated uptake, m e tabolism
, and excretion of benzene for m i ce, rats, and hum ans using experim e ntal data from the studies that were used to develop the m odel. However, the m odel is of lim ited value because it does not predict the kinetics of benzene

m e tabolites (Bois et al., 1991b). In addition, the concentration of benzene in fat is poorly predicted.

The m odel developed by Bois and Paxm an (1992) provided evidence that exposure rate had a strong influence on the rate of form ation of several im portant m e tabolites of benzene. This

m odel has three com ponents describing the pharm acokinetics of benzene and the form ation of

m e tabolites in the rat. Distribution and elim ination of benzene from a five-com partm e nt system

com posed of liver, bone m a rrow, fat, poorly perfused tissues, and well-perfused tissues m a ke up the first com ponent of the m odel. The bone m a rrow com p artm ent is included for its relevance to hum an leukem i a. Param e ter values for this com ponent were derived from the literature and from the previously published work of Rickert et al. (1979) and Sabourin et al. (1987). The second

com ponent describes the m e tabolic transform ations of benzene and its by-products in the liver and bone m a rrow. T he reactions are assum e d to follow Michaelis-Menten kinetics, except for the transform ation of benzene oxide into phenol, which occurs spontaneously and m a y be described as a first-order reaction. The third com ponent is the distribution of phenol. In addition to the com p artm ents described for benzene, phenol is also assum e d to distribute to the lung and gastrointestinal tract.

The m odel 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 m e tabolite production for m a le rats exposed to benzene (nose only) at three different concentrations and for

24three different exposure durations in com p arison with the experim e ntal data of Sabourin et al. (1989). The three exposure regim e ns were established to m a intain a constant concentration/tim e product. Sim u lation results indicated that the m odel m a y over- or underestim ate the level of urinary m etabolites.

More recent efforts on developm ent of the Bois and Paxm an m odel have focused on defining the physiological pharm acokinetic param e ter distributions needed to develop m odels useful in risk assessm ent (Spear and Bois 1994; Spear et al., 1991; W a tanabe and Bois, 1996; Bois et al., 1991a, 1996). Spear and Bois (1994) described the outcom e of their m odeling efforts to explain the basis for the paradoxical observation that although phenol is a m a jor initial

m e tabolite of benzene, a known carcinogen, a National Cancer Institute chronic study (NCI, 1980) did not dem onstrate carcinogenic activity for phenol. T he approach selected was to apply Monte Carlo m e thods using param e ter distributions coupled with a pass-fail fit criterion. The advantage of this approach is that it acknowledges that in m o st biological applications, there is no clear way to select a “best” set of fixed param e ters. O n the basis of the researchers’ m odeling effort, hydroquinone was rejected as the ultim ate 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 m e tabolism of phenol. Even though phenol is thought to be a key interm ediate in benzene m e tabolism leading to toxicity, orally adm i nistered phenol is neither carcinogenic or genotoxic (NCI, 1980). The authors found m a rkedly higher excretion of hydroquinone glucuronide after oral benzene exposure as com p ared with phenol. Also, phenol sulfate and phenol glucuronide excretion was m u ch lower following benzene exposure than following phenol exposure. This could be explained by differences in the zonal distribution of CYP2E1 and detoxification enzym e s 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 m o re likely to pass through to zone 3 and be oxidized to phenol.

W a tanabe and Bois (1996) exam ined three m e thods (m ultiplicative, additive, and allom e tric) to extrapolate physiological param e ter distributions across species, specifically from rats to hum ans. The results indicated that the m u ltiplicative and allom e tric techniques were able to extrapolate the m odel param eter distributions.

Bois et al. (1996) applied techniques from population pharm acokinetics, Bayesian statistical inference, and physiological m odeling to m odel distribution and m e tabolism in hum ans. Statistical distributions for the param e ters of a physiological m odel of benzene were derived on the basis of existing data. The relationship between the fraction of benzene

m e tabolized in bone m a rrow and benzene exposure was linear up to 10 ppm (32 m g /m 3 ). The

m e dian population estim ate of the fraction m e tabolized in bone m a rrow was 52% (90% confidence interval [CI] 47–67%). At levels approaching occupational inhalation exposure (continuous 1 ppm [3.2 m g /m 3 ]), the estim ated am ount m e tabolized in bone m a rrow ranged from 2 to 40 m g /day. H owever, this m odel has not been tested for its ability to predict data from other studies (Sm ith and Fanning, 1997).

25In sum m ary, PBPK m odels continue to im prove as additional data becom e available and are incorporated and as additional techniques from other scientific fields are applied to m odeling benzene dosim etry. However, the current published m odels are insufficiently refined to allow them to predict hum an m e tabolism accurately (Sm ith and Fanning, 1997). The key areas for refinem e nts appear to be the inclusion of the kinetics of the putative toxic m e tabolites of benzene or their stable precursors. If benzene m e tabolites such as hydroquinone/ benzoquinone, MUC, and/or benzene oxide are the toxic species, then PBPK m odels need to include descriptions of their kinetics if they are to be useful in im proving uncertainties in risk assessm ent.

3.6. TOXICOKINETICS SUMMARY

Benzene is alm o st com p letely absorbed from the gastrointestinal tract over a wide range of dose levels in rats and m i ce. In contrast, several hum an studies have indicated that the respiratory absorption of benzene is approxim a tely 50%. At a concentration of 10 ppm (31.9
mg / m 3 ), retention of benzene during a 6-hour exposure was 33% in rats and 50% in m i ce. Retention decreased at higher concentrations (S abourin et al., 1987). At high concentrations, benzene m e tabolism is saturated, and respiratory excretion of unm etabolized benzene increases. Derm al 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, interm ediate concentrations in blood, and the lowest concentrations in the Zym b al gland, nasal cavity, oral cavity, m a m m ary gland, and bone m a rrow. B enzene is preferentially stored in fat (Rickert et al., 1979). Metabolites of benzene are also found throughout the body, with phenol being elim inated from bone m a rrow m o re rapidly than catechol or hydroquinone (Rickert et al., 1979).

Despite extensive research, the m e tabolism of benzene to toxic m e tabolites is still not

com p letely understood. Metabolism of benzene by CYP2E1 is necessary for the expression of hem a totoxicity (Sam m e tt et al., 1979; Valentine et al., 1996). Following initial oxidation by CYP2E1 to benzene oxide, however, the benzene m e tabolic pathway branches to produce several putative toxic m e tabolites. Most research has focused on the hypothesis that phenol, catechol, and hydroquinone are produced in the liver and transported to the bone m a rrow, where hydroquinone is activated to p -benzoquinone by the action of MPO (Sm ith et al., 1989; Rushm ore et al., 1984; Low et al., 1995). However, the fact that adm i nistration of phenol does not produce the sam e effects as benzene is a problem in the phenolic m e tabolite hypothesis. A

com b ination of phenol and hydroquinone is required to reproduce the hem a totoxic effects of benzene (Eastm ond et al., 1987).

A ring-opening pathway of benzene m e tabolism leads to excretion of MA in the urine. The production of the highly reactive ring-opened dialdehyde MUC as an interm ediate in the pathway to MA has been dem onstrated in isolated hepatic m i crosom es; however, circulation of MUC in the blood has not been dem onstrated (W itz et al., 1996). Adm i nistration of MUC reproduces the hem a totoxic effects in m i ce; however, MUC is so reactive that it is not likely to reach the target tissues. The less reactive ring-opened m e tabolite CHO-M-OH has been detected in rat liver perfusate (Grotz et al., 1994) and has been shown to be hem a totoxic in m i ce (Zhang

26et al., 1995b). Thus, CHO-M-OH is another potentially toxic benzene m e tabolite, but there has been little research to define whether CHO-M-OH is quantitatively im portant.

Benzene oxide was believed to be too reactive to escape the liver, but it has recently been

m easured in the bloodstream of m i ce, and its dose to the body m a y exceed that of p

benzoquinone (Lovern et al., 1997; Lindstrom et al., 1997). Thus, several m e tabolites of benzene m a y play a role in inducing the toxic effects of benzene, and a com b ination of several

m e tabolites m a y be required to cause the full range of benzene-induced toxic responses. Large quantitative differences in the production of putative toxic benzene m e tabolites have been observed am ong different anim al species (Sabourin et al., 1989, 1992). Because few data exist on the proportion of different benzene m e tabolites produced in hum ans, there is considerable uncertainty in selecting the m o st appropriate anim al m odel for hum ans.

There is evidence to support several different m echanism s by which benzene m e tabolites

m a y cause hem a totoxicity. B enzene m e tabolites form adducts with both proteins and DNA (McDonald et al., 1994; Lutz and Schlatter, 1977). Adduct form ation with protein sulfhydryl groups inhibits tubulin polym erization during spindle form ation and the activity of topoisom e rase II in DNA replication and transcription (Irons and Neptun, 1980; Pfeifer and Irons, 1983; Frantz et al., 1996). These effects m a y explain the clastogenic activity of benzene. Benzene itself is not m u tagenic; however, m u tagenicity of several of its m e tabolites is well established (Dean, 1985). Benzene m e tabolites m a y also induce oxidative stress by depleting levels of GSH and by production of reactive oxygen species that react with cellular

m acrom olecules (Snyder et al. 1987; Brunm ar k and Cadenas, 1988; Rao and Snyder, 1995). Benzene also disrupts production of the cytokine IL-1, which is essential for hem a topoiesis. Niculescu et al. (1995, 1996) found that p -benzoquinone inhibits the activity of calpain, an
enzym e that catalyzes the processing of pre-IL-1 

###  to the m a ture cytokine. All of these effects

m a y be involved in causing the toxic effects of benzene, but it is not possible to determ ine which of these effects is the prim a ry m echanism for benzene toxicity.

At low exposure levels, benzene is excreted prim a rily in the urine as sulfate or glucuronide conjugatges of phenolic m e tabolites or as MA (McMahon and Birnbaum , 1991; Sabourin et al., 1987). At oral doses of less than 15 m g /kg, 14 C from benzene is nearly quantitatively recovered as m e tabolites in urine in rats and m i ce, but at a dose of 150 m g /kg, greater than 50% was cleared unm etabolized in the breath (Sabourin et al., 1987). Sim ilarly, less than 6% of 14 C from benzene was exhaled in the breath from inhalation exposure to 10 ppm (32
mg / m 3 ) in rats or m i ce, but at concentrations greater than 850 ppm (2718 m g /m 3 ) rats exhaled 48% and m i ce exhaled 14%. Urinary phenol and MA concentrations are correlated with benzene exposure level and can be used to m onitor benzene occupational exposure (Ghittori et al., 1993). However, the database for evaluating the relative im portance of different excretion pathways in hum ans is lim ited. Thus, there is a lim ited database for com p aring the im portance of hum an excretion pathways at varying dose levels with the results of experim e ntal anim al studies.

Several different PBPK m odels have been developed to m a them atically describe the uptake, distribution, m e tabolism
, and excretion of benzene (Sato, 1988; Medinsky et al., 1989a, b, c; Travis et al., 1990a, b; Bois and Paxm an, 1992). Each of these PBPK m odels has advantages, and som e successfully sim u late uptake, m e tabolism
, and excretion in m i ce, rats, and hum ans. However, the utility of PBPK m odeling to predict the dose of toxic m e tabolites to

27target organs in hum ans is lim ited by the incom p lete knowledge of the toxic m e tabolites and difficulties in identifying a suitable experim e ntal anim al m odel for hum ans.

4. HAZ ARD 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 victim s often were not known. However, lethal oral doses for hum ans have been estim ated at approxim a tely 125

m g /kg (Thienes and Haley, 1972). Accidental ingestion and attem p ted suicide with lethal oral doses of benzene have produced the following signs and sym p tom s : s taggering gait, vom iting, shallow and rapid pulse, som nolence, and loss of consciousness followed by delirium
, pneum onitis, collapse, and then central nervous system depression, com a , and death (Thienes and Haley, 1972). Lethality in hum ans has been attributed to respiratory arrest, central nervous system depression, or cardiac collapse (Greenburg, 1926). Ingestion of lethal doses m a y also result in visual disturbances or feelings of excitem e nt and euphoria, which m a y suddenly change to weariness, fatigue, sleepiness, convulsion, com a , and death (Von Oettingen, 1940).

4.1.2. Inhalation Exposure

4.1.2.1. Hematotoxicity

The recognition of benzene as etiologically significant in the developm ent of aplastic

anem ia, a potentially life-threatening suppression of bone m a rrow activities, and pancytopenia, a reduction of the cellular elem ents of the peripheral blood, dates back to the 19 th century (Goldstein, 1988). Exposure to benzene has also been associated with acute m y elogenous leukem i a (AML) and lym phoid m a lignancies in hum ans (Goldstein, 1988; Snyder, 1987), although the precise relationship between frank pancytopenia and the benzene-induced

m a lignancies rem a ins unclear (Snyder, 1987). That the hem a topoietic abnorm alities observed in hum ans exposed to benzene in an occupational setting have been reproduced in num erous anim al experim e nts points to their universality. The com p aratively low doses associated with their onset suggest that they m a y represent a sensitive effect of benzene toxicity.

Hum a n exposure to benzene occurs prim a rily via inhalation in the workplace, from gasoline vapors, tobacco sm oke, and autom o tive em issions. Individuals exposed to benzene exhibit bone m a rrow depression, as evidenced by anem ia (decreased RBC count), leukopenia (decreased W BC count), and/or throm bocytopenia (decreased platelet count). A depression of all three elem ents is called pancytopenia, and the sim u ltaneous depression of RBCs, W BCs, and platelets, accom p anied by necrosis of the bone m a rrow, is diagnostic of aplastic anem ia. Patients with aplastic anem ia also have exhibited m ild bilirubinem i a, changes in osm o tic fragility of erythrocytes, shortened erythrocyte survival tim e, increased fecal urobilinogen, and

m ild reticulocytosis (Aksoy, 1991). The bone m a rrow profile m a y vary from aplasia to hyperplasia, and these sym p tom s m a y vary in frequency and severity in different patients. Other hem a tologic changes observed in hum ans chronically exposed to benzene include decreased

28osm o tic resistance in leukocytes, decreased phagocytic function of neutrophils, reducedglycogen content, and decreased activity of peroxidase of neutrophils and increased deltaam inolevulinic acid activity in erythrocytes and increased coproporphyrins in the urine (Aksoy, 1991).

As sum m arized in Table 1, the large num ber of epidem iologic studies that have explored the health consequences of benzene exposure in the workplace provide a considerable body of inferential evidence for the com pound-related onset of hem a topoietic effects in chronically exposed subjects. It should be recognized that for m o st of the studies there are uncertainties and/or variabilities in precisely delineating the concentrations of benzene to which subjects were exposed. F or exam ple, uncertainty surrounding the past or present workplace concentrations of the chem ical m a y not perm it the dose to which subjects were exposed to be estim ated accurately. Even when an exposure term can be defined within an adequately cohesive range, its overall contribution to the understanding of the com pound’s dose-response characteristics will probably be lim ited to a single data point from any given study. Thus, the utility of such a study m i ght only arise through its consideration with others as a group, for exam ple, in a “m eta-analysis.” In such a survey, the establishm ent of a com b ined range for estim ated exposures and their toxicological consequences can perm it (1) the em ergence of a unifying concept of the

com pound’s toxicity to hum ans and (2) the delineation of a range of concentrations or doses within which the com pound’s toxicological threshold m a y reside.

The paragraphs that follow describe the m a ny epidem iologic studies that have helped to establish the onset of benzene’s hem a tologic effects.

In general, it is assum e d that for the occupational exposure studies under consideration, the levels of exposure to benzene were greater in earlier studies than in m o re recent studies, a reflection of tim es when occupational exposure standards and guidelines were less rigorous. For

exam ple, studies that have exam ined occupational exposure to benzene at relatively high concentrations include several that evaluated hem a tologic surveillance records of 459 rubber hydrochloride workers from the Pliofilm production departm e nts of a rubber products

29Table 1.Hematotoxicity of benzene—occupational exposurePopulation/IndustryNo. ofSubjectsExposureEffectsReferenceRotogravure printers677–3386 mg/m3benzenePancytopeniaErf and Rhoads, 1939Rotogravure printers,New York33235–3386 mg/m3benzene23 cases of significant cytopeniasGoldwater, 1941; Goldwaterand Tewksbury, 1941;Greenburg et al., 1939Shoe factory, Finland1471278 mg/m3benzene for ~10 yr107 cases of hematologic abnormalitiesSavilahti, 1956Shoe factory, Finland147Followup of Savilahti (1956) study, 9years laterPersistent cytopenias; one death from acuteleukemiaHernberg et al., 1966Rubber factory, Ohio11041597 mg/m3(mean, ~319 mg/m3)benzene83 cases of mild hematologic abnormalities, 25cases of more severe pancytopenia (9 requiredhospitalization, 3 of whom died)Wilson, 1942Watch industry216No exposure data; workers followed for10 years after cessation of exposure tobenzeneFour cases of persistentlydecreased blood counts;one death fromaplastic anemia 9 years aftercessation of exposureGuberan and Kocher, 1971Rubber-coatingindustry< 80–399 mg/m3benzene (prior toinstallation of control measures)Follow-up suggested mild persistent anemiaPagnotto et al., 1961Shoe manufacturing,TurkeyControls (healthyhospital workers)21710096–671 mg/m3benzene for 3 months–17years51 cases of hematologic abnormalities, including 6cases of pancytopeniaAksoyet al., 1971Various industriesusing benzenecontaining adhesives,Turkey32diagnosedwith blooddyscrasias479–1597 mg/m3benzene for 4months–15 years32 cases of significant aplastic anemia; eight deathsfrom thrombocytopenic hemorrhage and infectionAksoyet al., 1972Various industriesusing benzenecontaining adhesives,Turkey44diagnosedwith blooddyscrasiasFollow-up studyover 2–17 years after lastexposure to 479–2077 mg/m3benzene for4 months–15 yearsComplete remission in 23 (52%), death due tocomplications of pancytopenia in 14 (32%),leukemia in 6 (14%), myeloid metaplasia in 1 (2%)Aksoyand Erdem, 1978

30Table 1.Hematotoxicity of benzene—occupational exposure (continued)Population/IndustryNo. ofSubjectsExposureEffectsReferenceTurkish, various231Exposure to solvents, thinners, andsimilar materials in 40 small industries; inone facility, benzene concentrationsaveraged 351 mg/m3; in the remainder,benzene levels < 1 ppm351 mg/m3: two cases of acute myeloblasticleukemia in facilitywhere levels measured 351mg/m3< 1 ppm:ild abnormalities in 14 workers (6.1%);differential diagnosis of 186 workers, monocytosisin 8, eosinophilia in 5, and basophilia in 3; one caseeach of acute myeloblastic leukemia, Hodgkin’sdisease, and lymphomaAksoyet al., 1987Chemical productionfacility282 menJobs assigned to exposure categories:(1) < 6 mg/m3TWA(2) 6–29 mg/m3TWA(3) 32–77 mg/m3TWA(4) 80– ~96 ppm TWAexposures ranged from < 1 years to > 20yearsSlight statisticallysignificant (p<0.05) (but notbiologicallysignificant) decreases in bilirubin andRBC counts; no correlations between peripheralblood counts and latency,duration, or intensityofbenzene exposureTownsend et al., 1978Refinery303 menExposure to benzene levels of 1.70 mg/m3(mean) for an average of 8 years for whiteemployees and 4.5 years for nonwhiteemployees); average length of followup,13 yearsNo significant changes in hematologyof workers asa group; one case of multiple myeloma; one deathfrom multiple myeloma and one from malignantmelanomaTsai et al., 1983Chinese shoemakingand printing300Mean TWA exposures:Benzene—105 and 188 mg/m3(55 and 65months) for men and women,respectively;Toluene—173 and 154 mg/m3(95 and 73months);Mixture—45 mg/m3benzene + 68 mg/m3toluene and 45 mg/m3benzene + 79mg/m3toluene (159 and 120 months)Changes in absolute and/or relative peripheral cellcounts (statisticallysignificant atp<0.01 orp<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 (allgroups)Mixture—decreased lymphocyte counts (men;women; men and women combined); significanteosinophilia (men and women combined)Yin et al., 1987bm

31Table 1.Hematotoxicity of benzene—occupational exposure (continued)Population/IndustryNo. ofSubjectsExposureEffectsReferenceRubber industryWorkersUnexposed controlsWorkers in variousChinese industriesLow-exposure groupBelow medianAbove medianUnexposed controlsRubber industry459Mean estimated exposure from 1948 to1975, 48–64 mg/m3Significant increases inWBCs, RBCs, and Hgbduring the 1940’s (Pearson correlations:r = 0.50for WBCs, r = 0.44 for RBCs, and r = 0.71 forHgb), but not during the ensuing 25 years; rapiddecline in exposure for the workers consistent withincreases in blood count valuesKipen et al., 1988, 19892002680.03–4.5 mg/m38-hour TWA over a 10-year periodNo abnormal hematologyCollins et al., 199111222244Workers exposed to benzene, but not toother solvents, at TWA of 24 (lowexposure), 44 (below median), or 294(above median) mg/m3for an average of6.3 years; doses were measured bypersonal dosimetryfor 2 weeksSignificant increase in mean corpuscular volumeand significant decreasesin ALC, WBCs, RBCs,HCT, and platelets in above-median group;significant reductions in ALC, RBCs, and plateletcount in below-median group; but onlyreduction inALC was significant in low-exposure groupRothman et al., 1996a657Estimatesof benzene exposure (109mg/m3) for last 30, 90, or 180 daysbeforeblood tests were correlated withhematologic dataWeak correlation between RBC and benzeneexposure, but strong correlation between low WBCcounts and benzene exposure, with no evidence of athresholdWard et al., 1996

32m a nufacturer in Ohio (e.g., Kipen et al., 1988, 1989; W a rd et al., 1996). The data indicate that during the tim e period 1940–1948, significant increases in erythrocyte, Hgb, and leukocyte levels were observed am ong the workers, with the responses displaying a positive tim e-related trend (1940–1948). Thus, overall cell num bers and values appear to have been higher at the later tim e points, the prevailing levels of exposure to benzene becom i ng reduced as industrial hygiene practices im proved. Kipen et al. (1988, 1989) com p ared the utility of two exposure assessm ent

m odels for correlating the observed hem a tologic consequences with plausible estim ates of benzene exposure in the absence of definitive m onitoring data at the earlier tim e points. Thus, between 1948 and 1975, when the workers were exposed to m ean 8-hour TW A concentrations of only 15–20 ppm (48–64 m g /m 3 ), no significant relationships between benzene levels and any hem a tologic param e ters were observed. By contrast, significantly lower cell counts and hem oglobin concentrations were observed in exposed subjects, where the levels of benzene exposure were thought to be as high as 75 ppm (240 m g /m 3 ) at the earlier tim e points (pre-1948) (Kipen et al., 1989).

Cody et al. (1993) observed significant hem a tologic effects, including reduced RBC and W BC counts in 161 m a le rubber workers exposed to m e dian peak concentrations (i.e., only the peak concentrations for any given exposure tim e were reported) of 30–54 ppm or m o re for a 12-

m onth period during 1948. The 30 ppm value was considered a 1-year LOAEL for hem a tologic effects. In this rubber plant, workers who had blood dyscrasias were excluded from working in the high-benzene units. Furtherm ore, individual workers who had m o re than a 25% decrease in W BC counts from their pre-em p loym ent background count were rem oved 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 m e dian values (30–54 ppm ) reported by Crum p and Allen (1984) and used in the Kipen et al. (1988) and Cody et al. (1993) studies. Th ese are the sam e Pliofilm workers studied by Rinsky et al (1981, 1987).

The analysis of W a rd et al. (1996) differs from the earlier investigations of the sam erubber 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 W BC or RBC count, and m a tched controls were chosen from those tested within 6 m onths of the case’s blood test date. In contrast to the earlier analyses of hem a tologic screening data from the sam e plant, the W a rd et al. (1996) study used the entire data set, evaluated the exposure-response relationship on the basis of individual dose m e trics, and controlled for the temporal trends in preem ploym ent blood count screening.

The W a rd et al. (1996) analysis used hem a tologic screening data for 657 of 1037 individuals em ployed 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 estim ated benzene exposures using the job exposure m a trix 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 cum u lative exposure until the blood test date, were exam ined using conditional logistic regression. For W BC count there was a strong exposure response, and all of the exposure m e trics selected showed a significant relationship with low blood count. For RBC count there was a weak positive exposure response that was

33significant ( p <0.03) for one of the dose m e trics. The m a xim u m daily benzene exposure estim ate in this study was 34 ppm (109 m g /m 3 ). There was no evidence for a threshold for the hem a tologic effects of benzene exposure, suggesting that even exposure to relatively low levels (e.g., < 5 ppm [16 m g /m 3 ]) could result in hem a tologic suppression. The results of this study are consistent with test anim al studies that have dem onstrated a decrease in peripheral lym phocyte counts at benzene exposures as low as 10 ppm (32 m g /m 3 ) and a stronger effect of benzene exposure on lym phocytes than on red cells; therefore, if one assum e s that W a rd’s Pliofilm workers exposure estim ates are correct, they suggest that hem a tologic effects of benzene occur even at 5 ppm

. B ecause of controversy surrounding the exposure estim ates used by W a rd et al., this finding could be biased.

The nature of the problem is that the estim ates of early exposure to benzene (prior to 1946) are not based on any m easured am bient air inhalation data for any of the m o re than 1200 Pliofilm workers who begn work in the plants prior to 1946. Estim ates of exposure to benzene in the set of exposure indices chosen by W a rd and her associates (1996) were based on the exposure scenario described by Rinsky et al. (1981), which basically assum e d that the exposures were the sam e as the local and State standards established for benzene at the tim e (in the years prior to 1946). This assum p tion has been challenged by others (Crum p, 1992; Paustenbach et al., 1992), who claim that the exposures during this tim e were m u ch higher. Thus, the

assum p tion that the exposures could have been as low as < 5 ppm m a y not be correct. The weaknesses in the W a rd 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 exam ined the health effects of benzene, Fishbeck et al. (1978) studied the hem atologic characteristics of 10 Dow Chem i cal em ployees who had been exposed to high concentrations of the chem ical 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 TW A benzene concentrations exceeding 25 ppm (80 m g /m 3 ) for at least a part of the period of em ploym ent. 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 com pound, with an 8-hour TW A of 30–35 ppm (96–112 m g /m 3 ) and occasional “spikes” of up to 937 ppm (2990 m g /m 3 ). However, a 1963 study described by Fishbeck et al. (1992) revealed 8-hour TW As ranging from 37 to 132 ppm (118 to 422 m g /m 3 ), with four

em ployees’ exposure levels exceeding 100 ppm TW A (319 m g /m 3 ). Concurrent hem a tologic evaluations undergone by these em ployees included Hgb, HCT, RBC and W BC counts, sedim e ntation rates, platelet counts, differential blood counts, clot retention determ inations, and blood indices.

Key findings included enlarged RBCs, transient anem ia, reduced hem oglobin concentrations, and m ean corpuscular volum es (MCVs) exceeding m e dically accepted values. However, the authors noted that since 1963, the em ployees’ MCVs had shown a reduction from the highest values obtained, such that 5 of the 10 em ployees had MCVs within the norm al range. From these data, the authors concluded that at least som e of benzene’s effects on the hem a topoietic system m a y be reversible. This conclusion m a y be prem a ture. T his sm all group of 10 em ployees was a survivor population and therefore m a y have introduced a potential bias. They were selected because they were listed on current em ploym ent rolls of the com p any when the study was in progress. Em ployees who had left em ploym ent earlier, perhaps with a benzene34related 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 hem a tologic abnorm alities in 217 apparently healthy m e n exposed to benzene in the shoe m a nufacturing industry. The test battery included RBC, W BC, platelet, and differential cell count; packed cell volum e (PCV); bone m a rrow exam ination; and Hgb analysis. The subjects, 12–58 years old, were com p ared with 100 unexposed, age-m a tched hospital workers. Benzene levels ranged from 15 ppm (48 m g /m 3 ) during nonworking hours to 210 ppm (671 m g /m 3 ) during the use of adhesives containing benzene; exposure duration ranged from 3 m onths to 17 years. A total of 217 exposed workers were involved in the study, with 51 (23.5%) displaying hem a tologic abnorm alities. Of these, leukopenia was present in 9.7%, throm bocytopenia in 1.8%, leukopenia associated with throm bocytopenia in 4.6%, pancytopenia in 2.8%, acquired pseudo-Pelger Huet anom aly in 0.5%, eosinophilia in 2.4%, and basophilia in 0.5%.

Bone m a rrow exam ination of 11 of the 51 workers with abnorm al peripheral blood cell counts revealed (1) norm ocellularity with no abnorm al histology (two patients), (2) slight hypocellularity (three patients), (3) hypercellularity (one patient), or (4) norm ocellularity with abnorm al histology. Abnorm al findings present in various com b inations included m a turation arrest in the erythroid and m y eloid series, m a turation arrest in the granulocytic series, and

m a rked vacuolization in the m y eloid series. One worker with norm al hem a tology developed acute erythroleukem i a 4 years later. The investigators concluded, based on the incidences of the various blood abnorm alities, 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, m a inly from infection and bleeding.

In an industrial hygiene survey of 282 Dow Chem i cal workers who were categorized by job classification into four groups on the basis of their exposure to benzene, exposed individuals

were m onitored for a num ber of clinical chem istry and hem a tologic param e ters and com p ared with an age- and sex-m atched cohort of unexposed Dow em ployees 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 m g /m 3 ). However, as noted by the authors, neither the RBC counts nor any other hem a tologic param e ter correlated with exposure intensity, duration of exposure, or estim ated career dosage (Townsend et al., 1978). In fact, no consistent relationship between any m onitored param e ter and benzene exposure was apparent from their data.

A report by Yardley-Jones et al. (1988, 1990) described a survey in which blood sam p les were obtained from 66 refinery workers exposed to benzene at com p arative low concentrations and 33 unexposed em ployees serving as controls. Exposure levels of < 10 ppm (32 m g /m 3 ) were estim ated through consideration of the exposed em ployee’s workstation, supplem ented by a

sm all am ount of personal m onitoring data. Exposed subjects were assessed for the appearance of sister chrom atid exchange in peripheral lym phocytes and the ability of a subject’s 

###  -glucuronidase/sulfatase-treated urine to prom ote g ene reversion in the Am es test. In addition,

som e hem a tologic param e ters were m easured in whole blood, although the largely negative

35findings gave no indication of a cause-effect relationship for benzene within the estim ated
exposure range.

Bogadi-Sare et al. (2000), in a m e dical surveilance study, exam ined 49 fem a le workers in the shoem a king industry who were exposed to solvent m i xtures 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 confirm ed solvent exposure. The num ber of B-lym phocytes was lower in the shoe workers than in the controls ( p =0.01). Significant correlations were found between the level of im m unoglobulin G and benzene in the work atm o sphere. Confounding factors appeared to play no role in the im m unological findings. The authors found that exposure to benzene concentrations lower than 15 ppm can induce depression of the circulating B-lym phocytes.

Khuder et al. (1999) reported on a group of 105 petroleum workers exposed to 0.14–2.08
ppm (0.45–6.6 m g /m 3 ) benzene who, over tim e, had sm all but statistically significant falls in certain blood counts. However, there were problem s 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 com p are five hem a tologic
param e ters in 200 benzene-exposed workers with those of 268 unexposed em ployees in the sam e plant. The type of work is not identified in the study. Estim ates of the 8-hour TW A benzene concentration were obtained from m onitoring inform ation coupled with knowledge of the applicable work assignm ent and professional judgm ent. Benzene concentrations ranged from 0.01 to 1.4 ppm (0.03–4.5 m g /m 3 ) for exposed subjects. Param e ters under investigation included RBCs, W BCs, Hgb concentration, platelets, and MCV. The authors noted statistically significant differences in som e of the hem a tologic param e ters, based on dem ographic and lifestyle factors. Thus, when m u ltiple regression analyses were applied using the confounding factors and current exposure as independent variables, no significant correlations between benzene exposure and any hem a tologic param e ter was evident. Again, this was a study of currently em ployed workers and, as such, represented a healthy group of individuals, who, having survived potentially adverse effects from exposure to benzene and other environm ental 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) exam ined m o rtality patterns in a cohort of 454 form er and current m a le

em ployees who were selected by industrial hygienists from a larger group of 20,000 present and form er refinery em ployees. T he cohort m e m b ers were chosen because they had worked directly in the benzene, ethylene, arom atic distillate hydrogenation, or caim a n units, the principal petrochem i cal 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

sam e units were also included. Irregularly assigned workers to these units were not included, although they m a y have been exposed to benzene. Personal exposures to benzene were determ ined from 1979 to 1982. The m e dian air concentration was 0.53 ppmin the “benzenerelated units,” while in the refinery as a whole, the m e dian was 0.14 ppm

. In nonbenzene exposure areas, the m e dian was 0.07 ppm

. T his one-tim e “snapshot” of benzene exposure levels says little about what the exposures m a y have been in the past or about study duration, nor do

36they tell us whether these personal sam p les relate to the 303 m e dically surveyed workers discussed below, as they are a different group from the 454 who were part of the m o rtality cohort. F rom this study it is im possible to determ ine the concentrations of benzene to which the

m e dically surveyed workers were exposed.

The analysis of overall m o rtality in the m o rtality cohort of 454 revealed no excesses. In fact, m o rtality from all causes and from diseases of the circulatory system was significantly below expected values, based on com p arable groups of U.S. m a les. The authors concluded that the presence of a healthy-worker effect produced significantly lower risk estim ates, all causes (standard m o rtality 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

com p arison group of 823 people consisting of 10% of the workers who were em ployed in the

sam e 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, hem a tologic param e ters were analyzed in a separate group of 303 m a le “benzene workers” (according to the supervisors who selected them ) beginning in 1959 and ending in 1979. During this tim e, 11 deaths occurred in this group, whereas 21.5 were expected. Up to four hem a tologic tests per year were conducted on just these workers. Total and differential W BC counts, Hgb, HCT, RBC counts, platelets, and clotting tim es were calculated, and a 95% range and standard deviation were derived for each hem a tologic test. Unfortunately, as the authors pointed out, the data could not be com p ared with any “healthy population” standard, because existing “norm al” 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 sim ilar 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 em ployed group of workers (except for the 11 who died after the m e dical surveillance was begun) and, as such, a survivor population in which few or no adverse health effects would have been found. In addition, these m e dically surveyed em ployees were healthy and young. They had been exposed for a short tim e to m i nim a l 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 hem a tologic responses that are etiologically linked to exposure to benzene is evident, although the precise threshold level for these effects rem a ins uncertain. However, som e clarification of this issue has been provided by an ongoing collaboration between the National Cancer Institute and the Chinese Academ y of Preventive Medicine that has generated a series of reports on both large and sm all cohorts of subjects who were occupationally exposed to benzene.

Data from China describe clinical aplastic anem ia in factories with exposure levels said to range from 93 to 1156 m g /m 3

. Y in et al. (1987b) found 2676 cases of benzene

37poisoning—defined as a W BC count less than 4000 cells /m m 3 and a history of benzene exposure—from a review of m o re than 500,000 benzene-exposed workers in China. The
geom etric m ean concentration in 50,255 workplaces was 18.1 m g /m 3

. I t was reported that 64.6% of the workplaces had benzene exposure levels less than 40 m g /m 3

. F rom their review, the authors concluded that cases of benzene poisoning can occur even in factories with less than 40
mg / m 3 concentration of benzene.

Dosem eci et al. (1996) discussed the findings of a large-scale epidem iologic study involving subjects em ployed between 1949 and 1987 in 672 factories in 12 cities in China. The authors sought to establish plausible estim ates of the levels of benzene exposure through an industrial hygiene-based survey of all factory/work unit/job title/calender period data, providing

assessm ents that were subsequently correlated with one of six hypothetical exposure ranges between < 1 and > 50 ppm

. T hirty-eight percent of a total of 18,435 exposure estim ates were confirm ed by workstation m onitoring. The authors set hem a tologic criteria that defined benzene poisoning to include a W BC count of < 4000 cells/m m 3 , platelets < 80,000/m m 3 , etc., thereby allowing the identification of a case-defined subset of individuals who could be m a tched with estim ated levels of benzene exposure as described above. Relative risks of benzene poisoning by intensity of exposure during the m o st recent 1.5-year exposure period com p ared with those exposed to < 5 ppm (16 m g /m 3 ) yielded estim ated 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 m g /m 3 ), 20–39
ppm (64–125 m g /m 3 ), and > 40 ppm (> 128 m g /m 3 ), respectively. Clearly, if a significantly elevated risk of benzene poisoning is an indication of hem a totoxicity, then certainly exposures to benzene at 5–19 ppm are hem a totoxic. The decrease in ALC contributed substantially to the decrease in W BC count, based on the results of the Rothm an et al. (1996a) study discussed in the next review. In general, recent exposure history appeared to be m o re closely related to elevated risk than exposure duration, suggesting the com p aratively rapid onset of benzene hem a totoxicity. These results support a threshold of benzene hem a totoxicity in hum ans in the 5–19 ppm range, in broad agreem ent with the em erging exposure-response range that is apparent from the epidem iologic studies described in this section.

Som e data that have shed further light on this issue have been provided by the sam e group (Rothm an et al., 1996a) that conducted a sm all cross-sectional study that com p ared 44 workers exposed to a range of benzene concentrations sim ilar to those of Dosem eci et al. (1996) with 44 age- and gender-m atched unexposed controls. The purpose of this study was to show that exposure to benzene affects all the m a jor peripheral blood elem ents, with the ALC being the

m o st sensitive indicator of benzene-induced hem a totoxicity. F rom 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 sam e geographic area that did not use benzene were selected as test subjects and controls, respectively. The three workplaces using benzene included a factory that m a nufactured rubber padding for printing presses, a factory that

m a nufactured adhesive tape, and a factory that used benzene-based paint. The control workplaces included a factory that m a nufactured sewing m achines and an adm i nistrative facility. Subjects who had a history of cancer, therapeutic radiation, chem otherapy, or current pregnancy were excluded. Requirem e nts for inclusion in the study were current em ploym ent for at least 6

m onths in a factory that used benzene, m i nim a l exposure to other arom atic solvents, and no exposure to other known bone m a rrow-toxic chem icals or ionizing radiation. Controls who had

38no history of occupational exposure to benzene or other bone m a rrow-toxic agents were frequency-m atched to the exposed subjects on age (5-year intervals) and gender.

Benzene exposure was m onitored by organic vapor passive dosim etry badges worn by each worker for a full workshift on 5 days within a 1- to 2-week period prior to collection of blood sam p les. Benzene exposures of controls in the sewing m achine factory were m onitored for 1 day, but no exposure m onitoring was perform ed in the adm i nistrative facility. Com pound exposure was also evaluated by analyzing for benzene m e tabolites in urine sam p les collected at the end of the benzene-exposure period for the exposed subjects. Median levels of the urine

m e tabolites phenol, catechol, MA, and hydroquinone correlated positively with the 8-hour TW A personal air level m easurem e nts of benzene by exposure category (Table 2).

It is possible that air exposure to benzene m a y have been overestim ated am ong workers wearing m a rginally effective respirators. However, a com p arison of urinary phenol levels to the air benzene levels can be m a de. Urinary phenol is thought to increase in a linear fashion up to 100 ppm benzene. Roughly, 50 μg/m g creatinine is thought to be equivalent to an 8-hour TW A of 10 ppm benzene. For 19 workers exposed to m ore than 31 ppm who had an air and urine

sam p le taken on the sam e day, the m e dian benzene exposure was 73 ppm and the m e dian urine phenol level was 351 μg/m g creatinine, about what one would predict (letter from Nathaniel Rothm an, 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 sam p ling effort. Dosem eci et al. (1996) estim ated that only about 5% of m o re 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 hem a totoxicity occurred to that 1 individual in 19 who was reported to have a TW A of 328.5 ppm

. It is not always possible to obtain a perfect correlation between a surrogate m easure of exposure and an effect.

Historical benzene exposure of the subjects was evaluated by exam ining em ploym ent histories. However, this inform ation was not used to classify subjects by category of current
exposure. Data on age, gender, current and lifelong tobacco use, alcohol consum ption, m e dical history, and occupational history were collected by interview.

Six hem a tologic param e ters were evaluated: total W BC count, ALC, HCT, RBC count, platelet count, and MCV. Total W BC counts and ALCs were perform ed using a Coulter T540 blood counter. Abnorm al counts were confirm ed by hand. The Coulter com p any estim ated that the m achine used in this study counts about 10,000 W BCs to generate a total W BC count and the percent lym phocytes, which are m u ltiplied to provide the ALC. The hand count involves counting a standard 100 cells. The m a nual count also used to generate the percent lym phocytes in this study was the standard 100 cells. Much higher-quality data would be expected for the form er and less precision for the latter, due to sam p ling variation (as well as hum an scoring error). However, the Coulter counting m achine actually generated larger and m o re significant

39Table 2. Median benz ene urine metabolites, by exposure category, in a study of w o rkers exposed to benz ene in Shanghai, China, 1992 a

Exposure Category P henol 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/m g creatinine for all m e tabolites.

Source: Adapted from Rothm an et al., 1996a.

differences in the ALC between the exposure groups than did m a nual counts. The authors reported that previous studies that reported ALC and percent lym phocytes based on m a nual
counts m a y have underestim ated group differences due to nondifferential m easurem e nt error. Benzene m e tabolites in urine were m easured by an isotope dilution gas chrom atography/m a ss spectrom etry assay. Correlation analyses were com p leted using the Spearm a n rank order correlation m e thod. The W ilcoxon rank sum test was used to test for hem a tologic differences. Twenty-one of the 44 subjects in both the exposed and control groups were fem a le.

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 m e dian 8-hour TW A benzene exposure concentration for the 44 exposed workers was 31 ppm (99 m g /m 3 ). The 8-hour TW A for each subject was derived from the geom etric m ean of the five air m easurem e nts that were gathered on each of the exposed workers on the five different days during which the sam p les were taken. Exposure to toluene and xylene was less than 0.2 ppm (0.6 m g /m 3 ) 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 m e dian concentration of 31 ppm versus those exposed to less than the

m e dian concentration according to the 8-hour TW A. The m e dian (range) 8-hour TW A exposure concentration was 13.6 (1.6–30.6) ppm (43.4 [5.1–97.8] m g /m 3 ) for the low-exposure group and 91.9 (31.5–328.5) ppm (294 [101–1049] m g /m 3 ) for the high-exposure group. A subgroup of the low-exposure group com posed of 11 individuals, none of whose five separate m easurem e nts exceeded 31 ppm (100 m g /m 3 ) at any tim e during the m onitoring period, was also analyzed for ALC. The m e dian (range) 8-hour TW A exposure of these individuals was 7.6 (1–20) ppm (24 [3.2–64] m g/m 3 ). All five air m easurem e nts were com p leted within 1 to 2 weeks of the phlebotom ies on all subjects.

All six blood elem ents m easured were significantly different in the highest benzene exposure group when com p ared with controls (Table 3). ALC, W BCs, RBCs, HCT, and platelets were all significantly decreased, and MCV was significantly increased with increasing benzene exposure. These effects are consistent with the hem a totoxic effects of benzene shown in Aksoy (1989) and Goldstein (1988). ALC was reduced from 1.9 × 10 3 /μL blood in controls to

40Table 3. Comparison of mean peripheral blood counts (w ith standard deviations), by exposure status, in a study of w o rkers exposed to benz ene in Shanghai, China, 1992

Exposure Category WBC Count (×10 3 /  L blood) ALC (×10 3 /μL blood) RBC Count (×10 6 /  L blood)

4.7 (0.6)

4.6 (0.6) b

4.2 (0.6) c HCT (% ) MCV (  m 3 ) Platelets (×10 3 /  L blood)

Control (m edian = 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 Rothm an et al., 1996a.

1.6 × 10 3 /μL ( p <0.01) in the < 31 ppm (99 m g /m 3 ) group and to 1.3 × 10 3 /μL ( p <0.001) in the group exposed to > 31 ppm (99 m g /m 3 ) benzene. A dose-response relationship is evident in five

m easures of hem a totoxicity. The RBC and platelet counts were also significantly reduced in the < 31 ppm (99 m g /m 3 ) exposure group, but only ALC was significantly different in the lowestexposure subgroup (m edian = 7.6 ppm , 24 m g /m 3 ). In this subgroup of 11 workers, whose

m e dian 8-hour TW A exposure was 7.6 ppm (24 m g /m 3 ) benzene, ALC (1.6 × 10 3 /μL) was sim ilar to that of the larger group of 22 who were exposed to < 31 ppm benzene. Thus, ALC is the m o st sensitive m easure of benzene hem a totoxicity, representing a sentinel of a continuum of the likely hem a totoxic consequences of prolonged or enhanced exposure to benzene below 31
ppm

.

One feature of the Rothm an et al. (1996b) study was its attem p t to probe the m echanism s by which benzene m a y bring about the hem a totoxicological sym p tom s that were evident in the blood sam p les of exposed subjects. For exam ple, although m o st peripheral blood cell levels were decreased in exposed workers as com p ared with controls, peripheral cytokine levels (IL-3, IL-6, erythropoietin, granulocyte colony-stim ulating factor, and tissue necrosis factor-alpha) were sim ilar in exposed workers and controls, suggesting that benzene m a y not affect these growth factor levels in peripheral blood. By contrast, when the rates of m u tations were

m easured at the glycophorin A (GPA) locus, an increased frequency of som a tic cells with the NN variant was interpreted as an indication of the compound’s capacity to produce geneduplicating but not gene-inactivating m u tations at the GPA locus in bone m a rrow cells (Rothm an et al., 1995).

41The study by Rothm an et al. (1996a) is one of few occupational studies in which individual- and workstation-m onitored benzene concentrations provide a range of values that

encom p ass the exposure-response area into which, as judged by the data from other larger-scale epidem iologic studies, the point-of-departure for the com pound’s hem a tologic effects appears likely to fall. Further m a rking the study’s relevance as a basis for developing quantitative reference values for the com pound is the exposure-response relationship that em erged between ALCs and benzene concentrations. This observation is based on the inverse correlations between these param e ters 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 m e dian benzene concentration of 31 ppm (99 m g /m 3 ).

Further assurance that the study m e t 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 harm ful solvents was m i nim a l in the exposed group and that the study protocol fulfilled adequate criteria for age- and sex-m atching of exposed and control subjects.

As discussed above, a num ber of subgroups of benzene-exposed workers have been defined in the Rothm an et al. (1996a) study, including (1) a high-dose subset of 22 workers, the

m eans of whose 5-day passive dosim etry m easurem e nts for benzene were greater than the

m e dian (31 ppm [99 m g /m 3 ]) of the 5-day m ean benzene exposure concentration values for the benzene-exposed group as a whole; (2) a low-dose group whose 5-day m ean TW A benzene exposure values were lower than 31 ppm (99 m g /m 3 ) (m edian benzene concentration = 13.6 ppm [43.4 m g /m 3 ]); and (3) a subset of the low-dose group whose exposure levels in all 5-day consecutive breathing zone m easurem e nts were lower than 31 ppm (99 m g /m 3 ) (m edian benzene concentration = 7.6 ppm [24.3 m g /m 3 ]). Because a reduction in ALC was the only significant hem a tological effect obtained for all of the benzene groups, reduction in ALC was designated as the critical effect, and the lowest exposure m e dian of 7.6 ppm was defined as the LOAEL. There is som e concern in selecting the 7.6 ppm (24.3 m g /m 3 ) concentration as the point of departure for RfC developm ent, because the subjects in question (N = 11) m a y provide an inadequate groupwise com p arison 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 m e dian 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 prim a ry quantitative analysis for developing the RfC and the RfD, however, uses BMD m odeling approaches to m odel the Rothm an 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 m e thodology because m o re of the exposureresponse data can be included in the analysis, and one is not restricted to selecting one of the
exposure group m e dians as a LOAEL. The LOAEL-based RfC and RfD estim ates are provided for com p arison with the BMD-based estim ates. RfC and RfD estim ates are also derived from applicable experim e ntal anim al studies for com p arison with the hum an-based estim ates (Sections 5.1.5 and 5.2.5, respectively).

424.1.2.2. Neurotoxicity

Hum a ns have displayed sym p tom s of neurotoxicity following acute inhalation of relatively high concentrations of benzene (Snyder, 1987). Benzene produces generalized
sym p tom s such as dizziness, headache, and vertigo at levels of 250–3000 ppm (799–9584
mg / m 3 ) (Brief et al., 1980), leading to drowsiness, trem or, delirium
, and loss of consciousness at 700–3000 ppm (2236–9584 m g /m 3 ) (ATSDR, 1997). Death m a y result from exposure to 20,000
ppm (63,894 m g /m 3 ) benzene for 5–10 m i nutes (Sandm eyer, 1981). Neurological signs and
sym p tom s are sim ilar for both lethal and nonlethal exposures to benzene and for exposure to

m

u ltiple solvents including benzene. These neurological sym p tom s are reversible upon rem oval 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 m g /m 3 ), whereas toluene vapors did not exceed the 5 m g /m 3 level. Subsequently, the air levels of both benzene and toluene did not exceed the 5 m g /m 3 level. Seventy-four of the exam ined workers com p lained of frequent headaches (usually at the end of the workday), becam e tired easily, had difficulty sleeping, and com p lained of m e m o ry loss. The lim itations 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 exam ined the neurological effects of chronic exposure to benzene. Herregods et al. (1984) diagnosed transverse m y elitis in a young m a n exposed daily to benzene while working in the warehouse of a wholesale supplier of drugs and chem icals. The duration of exposure was not given by the authors. The diagnosis of benzene poisoning was based on differential diagnosis, ruling out other possible cau ses of poisoning. The clinical diagnosis was based on the patient’s occupation, high urinary phenol levels (28 m g /L in contrast to the norm al level, < 10 m g /L), and the coincidental decrease in urinary phenol and im provem e nt of clinical
sym p tom s over about 6 m onths after cessation of exposure. The diagnosis of transverse m y elitis is consistent with an acute transection of the spinal cord affecting both the gray and white m a tter.

Baslo and Aksoy (1982) conducted neurological exam inations on eight patients who had a history of chronic exposure to benzene and who were diagnosed with aplastic anem ia (cases 1–6) or preleukem i a (cases 7 and 8). The seven m a les and one fem a le ranged in age from 19 to 51 years (average age, 35.3 years). Cases 1, 2, 4, 5, 7, and 8 worked in the shoem a king industry, case 3 was a leather worker, and case 6 was a whistle m a ker who dipped plastic m a terial into an open vessel of solution known to contain 88.42% benzene and 9.25% toluene. At tim es, the concentrations of benzene in the working environm ents reached 210 ppm (671 m g /m 3 ). Exposure duration for the six patients with aplastic anem ia ranged from 6 m onths to 8 years (m ean, 6 years), and the period between the cessation of exposure and the neurological

exam inations ranged from 1 to 96 m onths. The two patients who had preleukem i a (cases 7 and 8) were exposed for approxim a tely 15 and 10 years, respectively, and were not exposed during the 8 and 6 m onths preceding neurological exam ination. Neurological, electrom yographical (EMG), and m o tor conduction velocity exam inations were perform ed on all patients (with the exception of case 8, who did not have the EMG test). At least three different m u scles were tested in the EMG and m o tor conduction velocity exam inations. Sensory conduction velocities

were m easured in the upper and lower extrem ities of cases 1, 5, 6, 7, and 8.

43Four cases (1, 2, 3, and 8) had other neurological sym p tom s that included decreased sensory vibration, atrophy of the leg, and exaggerated deep tendon reflexes; the other four (cases 4, 5, 6, and 7) had norm al neurological test results. EMG tests revealed neurogenic involvem e nt in cases 1 and 2 but were norm al in cases 3–7. Five of the cases (1, 2, 3, 7, and 8) had abnorm al

m o tor conduction velocity tests, involving at least one of the nerves tested. These effects were characterized by decreased m o tor 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 am plitudes 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 abnorm alities 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 exam ple, cases 3 and 4 had m oderate to severe hem a tologic findings and aplastic anem ia, but no significant neurological findings. For these two cases, long periods of nonexposure (53 and 96 m onths, respectively) preceded neurological testing.

The study is based on a sm all num ber of patients, lacks exposure and control data, and does not rule out the possibility that the workers were exposed to chem icals other than benzene and toluene; therefore, a reliable LOAEL or NOAEL could not be determ ined. The clinical im pression indicated a predom inant involvem e nt of the white m a tter in the peripheral nervous system

. The reports sum m a rized in this section have obvious deficiencies, such as sm all num bers of patients under study and a lack of inform ation regarding the intensity or duration of benzene
exposure. Thus, a reliable quantitative neurotoxic risk assessm ent is im possible at present. Nevertheless, the occurrence of sym p tom s following exposure and the am elioration of effects
upon rem oval 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 hum ans are sum m a rized in Table 4.

Vara and Kinnunen (1946) exam ined 12 fem a le workers who had gynecological disorders attributed to benzene exposure: t hree 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 m e nopause” but found no evidence of

m e nopause; however, the sym p tom s of toxicity, consisting of hyperm enorrhea and hypom enorrhea, ovarian hypoplasia, sterility, degeneration of the ovary, and/or dysfunction of the ovary were m o re severe in the two subjects nearing m e nopause. Peripheral blood counts revealed distinct leukopenia in four subjects and decreased neutrophils and platelets in m o st

44Table 4.Reproductive toxicity of inhaled benzene in humansPopulationNo.ExposureEffectsReferenceFemale workers with30Overt symptoms suggested exposure to12 subjects had menstrual disorders (e.g., hypermenorrhea)Vara and Kinnunen,symptoms of benzenelevels greater than 3.2 mg/m3and/or ovarian hypoplasia; women nearing menopause most1946poisoningseverelyaffected; 4 had leukopenia, and the majorityhadFemale factoryworkerswith benzene-inducedovarian hypofunctionNDNDFemale gluing operators360Exposureto gasoline (a major source ofbenzene) and chlorinated hydrocarbonsvia skin and inhalation; benzene levels,< 5 mg/m3; 40% of chlorinatedhydrocarbon measurements exceededpermissible limits by1.2- to 2.4-fold;controls had no chemical exposureFemale solvent workers174(1) < 131 mg/m3benzene(2)131 mg/m3benzene(3) controlFemale workers inshoemaking industry223Subjects exposed to benzene and toluene(compared with 327 controls)reduced neutrophil and platelet countsWorkers exhibiting ovarian hypofunction had decreasedascorbic acid concentrations inthe blood, as compared withclinicallyhealthyfemale 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 werestatisticallysignificantPushkina et al., 1968Mukhametova andVozovaya, 1972Yin et al., 1987aHuang., 1991ND = no data.

45subjects. The investigators tentatively attributed the sparseness in m e nstruation 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. Com pared with clinically healthy fem a le workers in the sam e factory, the levels of ascorbic acid (changes in ascorbic acid m e tabolism were used as an index of toxicity) were reduced in the blood of the workers who had ovarian dysfunction (subjects, 0.36 ± 0.02 m g %; controls, 0.49 ± 0.03 m g %; p <0.001). The details of this study were sparse, but the results were supported by a study in rats that showed a sim ilar effect. The justification for the changes in the ascorbic acid as an index of toxicity was not well desribed.

A Russian study evaluated 360 wom e n exposed to gasoline (a source of exposure to benzene) and chlorinated hydrocarbons via inhalation and skin contact (Mukham e tova and Vozovaya, 1972). Benzene levels were < 1.6 ppm (5 m g /m 3 ); chlorinated hydrocarbon levels were 1.2–2.4 tim es higher than the perm issible lim its. Com pared with workers with no exposure to benzene, the subjects had increased incidences of m e nstrual disturbances. As exposure duration increased, so did the num ber of prem a ture interruptions of pregnancy, the percentage of cases where theplacental m e m b rane rupture during parturition was im peded, and the num ber of cases of intrauterine asphyxia of the fetus. This study, although dem onstrating exposure-related reproductive effects and fetal toxicity, fails to distinguish between the effects of benzene and chlorinated hydrocarbons.

Yin et al. (1987a) conducted a com p rehensive study of the toxicity of benzene in which 300 workers exposed to benzene, toluene, or a m i xture of the two were exam ined for subjective
sym p tom s and hem a tologic and biochem i cal effects. For 174 wom e n, the m ean TW A exposure to benzene was 59 ppm (188 m g /m 3 ) (exposure duration, 65 m onths). The wom e n were divided into low-exposure (1–40 ppm [3.2–128 m g /m 3 ]) and high-exposure (41–210 ppm [131–671
mg / m 3 ]) groups and were exam ined separately for subjective and pancytopenia-related subjective
sym p tom s

. T he benzene-exposed subjects reported an increase in the incidence of hyperm enorrhea, 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 hem a tologic abnorm ality noted was a m ild but statistically significant decrease in lym phocyte count ( p <0.05), this hyperm enorrhea was tentatively considered to be related to pancytopenia. These findings are strengthened by the large num ber of subjects, the m easured exposure and statistical evaluation, and sim ilar observations in other studies; however, the significance of subjective sym p tom s is debatable.

A Chinese report (Huang, 1991) described reproductive dysfunction in fem a le workers exposed to benzene and toluene in the leather shoem a king industry. The exposed group (223

wom e n) exhibited increased incidence in the rate of m e nstrual disorders as com p ared with the 327 controls. The incidence rates of “m ense-blood anom aly” and dysm e norrhea had a tendency to increase with duration of em ploym ent. The incidences of spontaneous abortion and gestosis (toxem i a) 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.

46The data presented above provide lim ited evidence regarding reproductive effects of benzene exposure in hum ans and involve concom itant exposure to other chem icals. These studies did not provide good exposure m onitoring data or quantitative dose-response inform ation.

4.1.2.4. Developmental Toxicity

The following discussion reviews the developm ental toxicity of benzene in hum ans, focusing on m o re recent reports. Available studies on the developm ental toxicity of benzene in hum ans are sum m arized in Table 5. There is no convincing evidence that benzene produces

m a lform ations in hum ans or test anim als; however, a few studies suggest that the chem ical induces adverse reproductive effects in hum ans and causes retarded fetal growth in test anim als,

m a nifested m a inly as decreased fetal weight and delayed ossification in the presence or absence of m a ternal toxicity (Chatburn et al., 1981; ATSDR, 1997).

Benzene crosses the hum an placenta and is present in cord blood in quantities equal to or greater than those in the m a ternal blood (Dowty et al., 1976). Sum m aries of developm ental toxicity of benzene in hum ans have been described in individual case reports (Forni et al., 1971; Holm berg, 1979; Bordarier et al., 1991) and occupational studies (Vara and Kinnunen, 1946; Pushkina et al., 1968; Mukham e tova 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 m u ltiple chem icals including benzene. W ith only two exceptions (Bordarier et al., 1991; Mukham e tova and Vozovaya, 1972), all adult exposures in the studies sum m arized in these sections were via inhalation.

Only one study evaluated fetal effects follo wing inhalation of benzene alone. A worker who was exposed to benzene during her entire pregnancy suffered from severe pancytopenia and exhibited increased chrom osom al aberrations, but there were no effects on the fetus (Forni et al., 1971).

Axelsson et al. (1984) evaluated by questionnaire the outcom e of pregnancy am ong personnel em ployed in laboratory work at the University of Gothenburg between 1968 and 1979. Of 745 wom e n who responded to the questionnaire, 556 had been pregnant (a total of 1160 pregnancies). T he pregnant wom e n 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 rem e m b ered using benzene during the first trim ester of pregnancy and 5 used phenol. A slightly increased but not significant difference in the m i scarriage rate was found (relative risk = 1.31, 95% CI = 0.89–1.91). There were no differences in perinatal death rates or prevalence of m a lform ations between infants whose m o thers were exposed to solvents and those who were not exposed. The investigators suggested that shift work, num ber of pregnancies, and age m a y have contributed to an increase in

m i scarriage rate in the group of pregnant wom e n not exposed to solvents, resulting in an underestim ation of the m i scarriage rate of the exposed wom e n. 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 m u tagenic effect occurring during laboratory work (laboratory work the

wom a n had stopped working at the laboratory, thereby altering the m i scarriage rate in the

47Table 5.Developmental toxicity of benzene—humansPopulationNo.ExposureEffectsReferencePregnant worker1Exposure to benzene during entireMaternal effects included severe pancytopenia and increasedForni et al., 1971chromosomal aberrations; no fetal effectspregnancyEmployees in universitylaboratory(~745subjects, 1160pregnancies), dividedinto those with andwithout exposure toorganic solvents745Responders to questionnaire reportedexposure to at least 14 solvents; 41workers remembered using benzeneduring the first trimester of pregnancyand 5 used phenolAll women exposed to solventshad slight but not statisticallysignificant difference in miscarriage rate over those not exposed(RR = 1.31, 95% CI = 0.89–1.91); 35 of 41 workers exposed tobenzene delivered, 1 had induced abortion, and 5 miscarried(miscarriage rate for benzene-exposed subjects, 12.2%; miscarriagerate for all responders to questionnaire, 11.1%; miscarriage rate forunexposed responders, 10.1%); all 5 workers exposed to phenoldelivered; exposure to solvents did not affect perinatal death rates orthe incidence of malformationsAxelsson et al.,1984Female gluing operators360Exposureto gasoline (a major source ofbenzene) and chlorinated hydrocarbonsvia skin and inhalation; benzene levels,< 5 mg/m3; 40% of chlorinatedhydrocarbon measurements exceededpermissible limits by1.2- to 2.4-fold;controls had no chemical exposureSpontaneous abortions and premature births (17.2% vs. 4.9% incontrols), incidence of late membrane rupture, and intrauterineasphyxia of the fetus increased with exposure durationMukhametova andVozovaya, 1972Adult female workersand 14 of their children29Adults exposed to benzene and otherorganic solvents during pregnancy(compared with 42 control adults and 7 oftheir children)Lymphocytes from adults exhibited approximatelytwofold increaseover controls in incidence of chromosomal aberrations and breaks;their children exhibited increased frequencyof chromatid breaks,isochromatid breaks (p<0.01, 14 children), and sister chromatidexchanges (p<0.001, 4 children) in lymphocytesFunes-Cravioto etal., 197732-year-old pregnantworker1Personal interview revealed exposure tobenzene and other solvents(dichloromethane, methanol, and ether)in laboratoryduring first trimester ofpregnancy;compared with matchedcontrolStillborn anencephalic fetusHolmberg, 1979

48Table 5.Developmental toxicity of benzene—humans (continued)ReferenceBordarier et al.,1991EffectsFollowing normal delivery,infant exhibited slight dysmorphy(hypotelorism and deep nasal bridge), moderate axial hypotonia andabnormal ocular movements; at1.5 months of age, child wasmicrocephalic, had severe axial hypotonia, severe peripheralhypertonia, and bilateral optic atrophy, and CT scanning revealedbilateral porencephalic cavities that created communication betweenlateral ventricles and subarachnoid space; interventricular septumlacking; child died fromaspiration pneumonia at 2 months of ageExposure21 intramuscular injections of benzene inan unsuccessful attempt to induceabortion during first trimester ofpregnancyNo.1Population23-year-old femaleRR = relative risk.CT = computerized transverse tomography.

49control group. Another weakness in the study was that the subjects had worked in the laboratory during 1968–1979, and som e had to rem e m b er several years back.

Four other studies evaluated the developm ental toxicity of m u ltiple solvents. These effects consisted of spontaneous abortions, prem ature births, and effects on “condition of the fetus” (Mukham e tova and Vozovaya, 1972), chrom osom al abnorm alities (Funes-Cravioto et al., 1977), and stillbirth (Holm b erg, 1979).

Savitz et al. (1989) exam ined the effects of parental (paternal as well as m a ternal) occupational exposures on fetal developm ent. The subjects, em ployed in various industries, were exposed to a num ber of chem icals. Sum m arized here are the data pertaining only to benzene exposures. The investigators accessed the National Natality and Fetal Mortality surveys to obtain data on probability sam p les of live births and fetal deaths that occurred in the United States in 1980 am ong m a rried wom e n. These data were m e rged with data provided by m o thers from questionnaires and from inform ation from m e dical care providers for the public-use data
set. Om itted from the analysis were unm arried m o thers, m o thers who had not worked within 12

m onths of delivery, plural births, and births of unknown plurality.

The analysis, which was based on three case control studies of pregnancy outcom e , included case groups of stillbirths, preterm deliveries (birth before 37 weeks of gestation), and

sm all-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 em ploym ent and on past studies and review of industrial processes. Linkages of none, low, m e dium , high, or unknown were assigned to each agent, representing probability (rather than intensity) of exposure. Unexposed working wom e n served as controls. The odds ratios (ORs) relating exposure to effect were adjusted for selected potential confounders. Elevated ORs were found for m a ternal exposure to benzene and stillbirth (OR = 1.3, 95% CI = 1.0–1.8). Increased risks across low-, m e dium -, 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, m e dium , 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 lim itations. T hese included sm all population sizes, poor quality of exposure inform ation, and absence of statistical testing and the high rate of nonresponse to questionnaires

am ong wom e n who were under 20 years old, 40 or m o re years old, and black and who had parity four or greater, little or no prenatal care, and low education. However, in spite of these lim itations, the authors concluded that the results of the study encourage further evaluation of the developm ental effects of paternal exposure to benzene.

The data presented above provide inconclusive results regarding developm ental toxicity of benzene in hum ans. Most studies consisted of sm all num bers of subjects, lacked im portant experim e ntal details, and involved (in alm o st all cases) concom itant exposure to other chem icals, and they did not provide m onitoring data or quantitative dose-response inform ation.

4.2. ACUTE AND CHRONIC STUDIES IN EXPERIMENTAL ANIMALS

504.2.1. Hematotoxicity

4.2.1.1. Oral Exposure—Subchronic Studies

Subchronic oral studies have been conducted in F344 rats and B6C3F1 m i ce of both sexes (NTP, 1986; Huff et al., 1989) and fem a le W i star rats (W olf et al., 1956). Other subchronic oral studies have been conducted to exam ine im m unotoxicity, and these are described in Section 4.2.4.1.

Hsieh et al. (1988b) exposed m a le CD-1 m i ce to benzene in drinking water at concentration levels 0, 31, 166, or 790 m g /L (0, 8, 40, or 180 m g /kg/day) for 28 days. Doserelated hem a tologic effects (erythrocytopenia, leukocytopenia, lym phocytopenia, increased MCV) were observed at all exposure levels. It can be estim ated that these dose levels in drinking water are roughly equivalent to inhaled doses of 6.4, 32, and 154 ppm

. A m a jor drawback of this study is that there were only five anim als 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

m i ce (10/species/group/sex; 6–8 weeks of age) were treated with 0, 25, 50, 100, 200, 400, or 600

m g /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 m g /kg/day. An additional five anim als/species/group/sex were tested at the 0, 200, or 600 m g /kg dose levels and killed at 60 days of treatm e nt. H em atologic analyses were perform ed on all the anim als killed at 60 days and on five anim als/species/group/sex at the end of the study. In addition, necropsies were perform ed on all anim als, and the spleens of all anim als were exam ined histopathologically.

No com pound-related deaths were observed for rats. Final body weight depression of



###  10% relative to controls was observed in m a le and fem a le rats at dose levels of 200 m g /kg and greater. S ignificant ( p <0.05) leukopenia and lym phocytopenia were observed in m a le and fem a le rats after 60 days of treatm e nt with 200 or 600 m g /kg (the only treatm e nt groups tested on day 60). On day 120 of treatm e nt, significant leukopenia and lym phocytopenia were observed in fem a le rats at 25 m g /kg and higher, and significant lym phocytopenia was observed in m a le rats at 400 m g /kg. Lym phoid depletion of B-cells in the spleen was observed in 100% of

m a le and fem a le rats exposed to 600 m g /kg for 60 or 120 days. Increased extram edullary hem a topoiesis in the spleen was observed in four of five m a le and three of five fem a le rats treated with 600 m g /kg for 120 days. This study identified a LOAEL of 25 m g /kg (18

m g /kg/day) in fem a le rats and a LOAEL of 200 m g /kg (143 m g /kg/day) in m a le rats for hem a tologic effects following treatm e nt by gavage for 17 weeks. The LOAEL for fem a le rats was at the lowest dose tested; thus, no NOAEL was established.

There were no com pound-related deaths in the m i ce. A final body weight depression of 7% was seen in the 100 m g /kg dose group. Trem ors were observed interm ittently in m a le and fem a le m i ce treated with 400 or 600 m g /kg. No leukopenia or lym phocytopenia was observed in

m a le or fem a le m i ce after 60 days of treatm e nt with 200 or 600 m g /kg. At 120 days, significant ( p <0.05) leukopenia and lym phocytopenia were observed in m a le m i ce at dose levels of 50

m g /kg and greater and in fem a le m i ce at 400 (only lym phocytopenia) and 600 m g /kg. A

51NOAEL of 25 m g /kg (18 m g /kg/day) and a LOAEL of 50 m g/kg (36 m g/kg/day) for hem a tologic effects were identified in m a le m i ce treated by gavage for 17 weeks. A NOAEL of 200 m g /kg (143 m g /kg/day) and a LOAEL of 400 m g /kg (286 m g /kg/day) for hem a tologic effects were identified in fem a le m i ce treated by gavage for 17 weeks.

Fem a le W i star rats (10/group) were treated by gavage with benzene in olive oil 5 days/week for 6 m onths (W olf et al., 1956). The reported doses were 0, 1, 10, 50, or 100 m g /kg (0, 0.7, 7.1, 35.7, and 71.4 m g /kg/day). Param e ters m easured included m o rtality, clinical signs, body and organ weights, hem a tology, blood biochem i stry, bone m a rrow counts, and gross and

m i croscopic pathology of lungs, heart, liver, kidneys, spleen, testes, adrenals, and pancreas. Leukopenia was reported for 10 m g /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 m g /kg had “no evidence of ill effects.” This study identified a NOAEL of 0.7 m g/kg/day and a LOAEL of 7.1 m g/kg/day for hem a tologic effects in fem a le rats treated by gavage for 6 m onths.

Aoyam a (1986) showed that a 14-day exposure of m i ce to 50 ppm (162 m g /m 3 ) 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 m i ce 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, m a les (60/group) were adm i nistered doses of 0, 50, 100, or 200

m g /kg (0, 36, 71, or 143 m g /kg/day), and fem a les (60/group) were adm i nistered doses of 0, 25, 50, or 100 m g /kg (0, 18, 36, or 71 m g /kg/day). Survival decreased with increasing dose in rats of both sexes and was significantly decreased ( p <0.05) at 200 m g /kg in m a les and at 50 and 100

m g /kg in fem a les. Body weight depression of 

###  10% relative to controls was observed in m a le rats treated with 100 m g /kg. Dose-related leukopenia was significant in fem a le rats treated with



###  25 m g /kg for 3, 6, 9, and 12 m onths; leukocyte levels were com p arable with controls after 15, 18, 21, and 24 m onths of treatm e nt. In m a le rats, dose-related leukopenia was significant at 

###  50

m g /kg. Lym phoid depletion was observed in the thym us of 0/44, 4/42, 8/41, and 10/34 m a le rats treated with 0, 50, 100, and 200 m g /kg benzene, respectively. In the spleen, lym phoid depletion was observed in 0/49, 19/58, 8/47, and 23/47 m a le rats treated with 0, 50, 100, and 200 m g /kg, respectively. This study identified a LOAEL of 25 m g /kg (18 m g /kg/day) for leukopenia and lym phocytopenia in fem a le F344 rats and 50 m g /kg (36 m g /kg/day) in m a le F344 rats. These were the lowest doses tested, and thus no NOAEL was identified.

B6C3F1 m i ce (60/sex/group) were adm i nistered doses of 0, 25, 50, or 100 mg/kg (0, 18, 36, or 71 m g /kg/day). Survival decreased with increasing dose in m i ce of both sexes and was significantly decreased ( p <0.05) at 100 m g /kg. Body weight depression of 

###  10% relative to controls was observed in m i ce of both sexes treated with 100 m g /kg. Significantly decreased leukocyte counts were observed in m a les after 3, 6, 9, 12, 15, 18, and 21 m onths of treatm e nt with 50 or 100 m g /kg, but m a les treated with 25 m g /kg had significantly decreased leukocyte counts only after 6 and 21 m onths. In fem a le m i ce, leukopenia was observed only at 12 and 18

m onths, in both cases significant at all treatm e nt levels. Significantly decreased lym phocyte counts were observed in m a les after 3, 6, 9, 12, 15, 18, and 21 m onths with 50 or 100 m g /kg, but

52m a les treated with 25 m g /kg had significantly decreased lym phocyte counts only after 12

m onths. In fem a le m i ce, significant lym phocytopenia was observed at 

###  25 m g /kg at 12 and 18

m onths and at 100 m g /kg at 3 m onths. Hem a topoietic hyperplasia of the bone m a rrow was observed in 0/49, 11/48, 10/50, and 25/49 m a le m i ce and in 3/49, 14/45, 8/50, and 13/49 fem a le

m i ce treated with 0, 25, 50, or 100 m g /kg, respectively. Increased splenic hem a topoiesis was observed in 5/49, 9/48, 19/49, and 24/47 m a le m i ce and in 9/49, 10/45, 6/50, and 14/49 fem a le

m i ce treated with 0, 25, 50, or 100 m g /kg, respectively. In the fem a le m i ce, 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 com p ared with controls. This study identified a LOAEL of 25 m g /kg (18 m g /kg/day) for leukopenia and lym phocytopenia in m a le and fem a le B6C3F1 m i ce. 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 treatm e nt of rodents with benzene were perform ed at the Bologna Institute of Oncology, including 52- and 104-week studies on Sprague-Dawley and W i star rats and Swiss and Rf/J m i ce (Maltoni et al., 1983, 1985). Lim ited inform ation regarding noncarcinogenic effects were reported for Sprague-Dawley rats, but only carcinogenicity data were published for W i star rats and m i ce. No statistical inform ation was included, m a king 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 m g /kg benzene in oil 4–5 days/week for 52 weeks and then observed them until death. Expanded doses were 0, 32, and 161 m g /kg/day. In addition, Sprague-Dawley rats (7 weeks of age, 40–50/sex/group) were treated by gavage with 0 or 500 m g /kg benzene in oil 4–5 days/week for 104 weeks and then observed until death. The expanded doses were 0 and 321 m g /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 m o rtality in the first portion of the study was due to direct toxic effects of treatm e nt, and in the later portion it was partially due
to tum o rs. Mortality was sim ilar to that of controls during treatm e nts with 500 m g /kg for 92 weeks; body weight appeared to be som e what depressed relative to controls. In Sprague-Dawley rats exposed to 500 m g /kg for 84 or 92 weeks, decreased total RBCs (only at 92 weeks), W BCs, and lym phocytes were observed. Insufficient inform ation was provided to establish LOAEL or NOAEL levels from these studies.

4.2.1.3. Inhalation Exposure—Subchronic Studies

The following discussion sum m a rizes m o re recent studies of the effects of inhaled benzene on the peripheral blood of experim e ntal anim als. The available studies on the hem a totoxicity of inhaled benzene to test anim als are sum m a rized in Table 6. These studies form a database that supports the understanding that exposure to benzene at high concentrations causes hem a totoxic effects, including bone m a rrow suppression and the greater sensitivity of

m i ce over that of rats. Benzene-induced hem a totoxicity has been observed from short-term as well as long-term exposure to the chem ical.

53Table 6.Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposureStrain/SpeciesEndpointsExposureEffectsLOAEL/NOAEL(mg/m3)ReferenceCD-1 mice(11–12 males/group)Body weight, cells inperipheral blood and bonemarrow, spleen weight, totalnucleated cells per spleen,lymphocytes0, 32 mg/m3, 6hr/day, 5days/wk for10 weeksIncreased spleen weight, total nucleatedcells per spleen, and nucleated cells perspleenLOAEL: 32Green et al.,1981a, bCD-1 mice(30 males and30females/group)HCT; Hgb; WBC, RBC,platelet, reticulocyte, anddifferential counts; MCV;mean corpuscular Hgb;myeloid/erythroid ratio frombone marrow; leukocytealkaline phosphatase; RBCglycerol lysis time; and othermore general parameters ofsubchronic toxicity0, 3.2, 32, 96,or 958 mg/m36 hr/day, 5days/wk forup to 91 days;serialsacrifices after7, 14, 28, 56,and 91 days ofexposureAt 958 mg/m3, decreased HCT, totalHgb, RBC, leukocyte, and plateletcounts; decreased myeloid/erythroidratios and percentage of lymphocytes;increased MCV, mean cell Hgb,glycerol lysis time, and incidence andseverity of RBC morphological changes(statistically significant [p<0.05] formales and/or females); many effectsoccurred and were statisticallysignificant by 14 days, persistingthroughout exposure; microscopicchanges in the thymus, bone marrow,lymph nodes, spleen, ovaries, and testes;changes in males occurred more oftenand with greater severity than those inthe femalesLOAEL: 958NOAEL: 96Ward et al.,1985AKR/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, cytogeneticanalyses of bone marrow, andmicroscopic examination oflung, liver, spleen, and kidney958 mg/m3benzene, 6hr/day, 5days/wk forlifeEarly mortality (mean survival time, 11weeks for test animals, 39 weeks forcontrols), severe lymphocytopenia, bonemarrow hypoplasia, granulocytosis, andreticulocytosisLOAEL: NDNOAEL: NDSnyder et al.,1978

54Table 6.Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposure (continued)ReferenceSnyder et al.,1980Snyder et al.,1980LOAEL/NOAEL(mg/m3)LOAEL: NDNOAEL: NDLOAEL: NDNOAEL: NDEffectsNo statistically significant differencesbetween test and control mice in mediansurvival time, rate of weight gain orlymphoma type or occurrence;statistically significant lymphocytopenia(cell counts were depressed to ~65% ofnormal throughout exposure); depressedRBC counts (± 2 standard error of meanat 9/19 monitoring periods); increasedneutrophils (± 2 standard error of meanat 3/19 monitoring periods); bonemarrow hypoplasia in 20% of thetreated mice (p<0.05)Decreased survival (median time 41weeks, test animals; 75 weeks,controls); lymphocytopenia (WBCcounts, ~3000/mm3or 15% of normal);anemia (RBC counts, 8 × 106/mm3or~80% of normal); neutrophiliaappearing at 17 weeks; abnormal bloodcell morphology; neutrophilic shift toleft; bone marrow hyperplasia in 33% oftest animals, none in controls;hematopoietic neoplasmsExposure319 mg/m3benzene, 6hr/day, 5days/wk forlife958 mg/m3benzene, 6hr/day, 5days/wk forlifeEndpointsRBC and WBC counts(Coulter) and differentialcount; microscopicexamination of lung, liver,spleen, kidney, and bonemarrowRBC and WBC counts(Coulter) and differentialcount; microscopicexamination of lung, liver,spleen, kidney, and bonemarrowStrain/SpeciesAKR/J mice(50 males, test;50 males,control)C57BL/6 mice(40 males, test;40 males,control)

55Table 6.Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposure (continued)ReferenceSnyder et al.,1982Ward et al.,1985LOAEL/NOAEL(mg/m3)LOAEL: NDNOAEL: NDLOAEL: 958NOAEL: 96EffectsDecreased median survival time (25.5weeks for benzene-exposed animals,52.7 weeks for controls); depressedperipheral RBC and lymphocyte counts(± 2 standard error of mean);neutrophilia; abnormal blood cellmorphology; and a shift to immaturemyeloid cells (at 217 days); tumorincidence (5/40) not significant; acutemyeloblastic leukemia in one exposedanimal, chronic myelogenous leukemiain one; bone marrow hyperplasia in 9/35animals without neoplasia, bone marrowhypoplasia in 11, splenic hemosiderinpigments in 6, and splenic hyperplasiain 19Statistically significant (p<0.05)decrease in WBC counts (males on day14 and females on days 14–91) andslightly decreased femoral marrowcellularity at 958 mg/m3(300 ppm)Exposure958 mg/m3benzene, 6hr/day, 5days/wk forlife0, 3.2, 32, 96,or 958 mg/m36 hr/day, 5days/wk forup to 91 days;serialsacrifices after7, 14, 28, 56,and 91 days ofexposureEndpointsRBC and WBC counts(Coulter), and differentialcount; microscopicexamination of lung, liver,spleen, kidney, and bonemarrowHCT; Hgb; WBC, RBC,platelet, reticulocyte, anddifferential counts; MCV;mean corpuscular hemoglobin;myeloid/erythroid ratio frombone marrow; leukocytealkaline phosphatase;erythrocyte glycerol lysis time;and other more generalparameters of subchronictoxicityStrain/SpeciesCD-1 mice (40males, test; 40males, control)SpragueDawley rats(10 males and10females/group)

56Table 6.Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposure (continued)Strain/SpeciesEndpointsExposureEffectsLOAEL/NOAEL(mg/m3)ReferenceSpragueDawley rats(45 males, test;27 males,control)WBC and RBC counts(Coulter), reticulocyte counts,differential counts, Hgb, HCT,lactic acid dehydrogenase,RBC acetylcholinesterase, andreduced GSH, cytogeneticanalyses of bone marrow, andmicroscopic examination oflung, liver, spleen, and kidney958 mg/m3benzene, 6hr/day, 5days/wk forlifeMild weight depression,lymphocytopenia, trend to anemia, fattychanges in 77% of bone marrowsamplesLOAEL: NDNOAEL: NDSnyder et al.,1978Female BDF1mice (4/group)WBCs, reticuloctyes,differential counts, and assaysfor CFU-S, BFU-E, CFU-E,and GM-CFU-C0, 319, 958, or2875 mg/m3, 6hr/day, 5days/wk for16 weeksReduced lymphocyte count at319mg/m3, decreased CFU-E, BFU-E,CFU-S, and GM-CFU-CLOAEL: 319NOAEL: NASeidel et al.,1989MaleC57BL/6Jmice (5/group)Peripheral blood counts, spleenand bone marrow BFU-E andCFU-E32 mg/m3, 6hr/day, 5days/wk forup to 178 daysReduced circulating RBCs, lymphocytecounts, bone marrow CFU-E and BFUE, and splenic CFU-ELOAEL: 32Baarson et al.,1984Male andfemale C57/6BNL mice(5–10/group)WBC, RBC, differentialcounts, HCT, bone marrowcellularity, and CFU-S in bonemarrow and spleen0, 32, 80, or1278 mg/m3, 6hr/day, 5days/wk for 2weeksReduced lymphocyte counts after 2weeks of exposure at 80 mg/m3but noeffect at 32 mg/m3LOAEL: 80NOAEL: 32Cronkite et al.,1985Male SpragueDawley ratsErythrocyte and lymphocytecounts0 or 319mg/m3, 6hr/day, 5days/wk forlifeNo significant reductions in erythrocyteor lymphocyte countsNOAEL: 319AmericanPetroleumInstitute, 1983

57Table 6.Peripheral blood and hematopoietic effects of benzene in animals—inhalation exposure (continued)Strain/SpeciesEndpointsExposureEffectsLOAEL/NOAEL(mg/m3)ReferenceMale B6C3F1mice (24/groupexposed, bonemarrow testson3–16/group)Bone marrow; cell counts,CFU-E, GM-CFU-C, and Band T-lymphocytes0, 3.2, 32,319, or 639mg/m3, 6hr/day, 5days/wk for 1,2, 4, or 8weeksReduced bone marrow cellularity,progenitor cells, and differentiatinghematopoietic cells, at319 mg/m3, butno effects observed at32 mg/m3LOAEL: 319NOAEL: 32Farris et al.,1996, 1997aMale B6C3F1mice(3 or 9/group)Splenic, thymic, and femorallymphocytes and labelingindex of femoral lymphocytes0, 3.2, 16, 32,319, or 639mg/m3, 6hr/day, 5days/wk for 1,2, 4, or 8weeksPersistent and rapid reductions insplenic and femoral lymphocytes at319 mg/m3, transient reduction ofthymic cell count and splenic Blymphocytes at 32 mg/m3after 2 weeks,but comparable with controls by 4weeksLOAEL: 319NOAEL: 32Farris et al.,1997bMale SpragueDawley rats(16/group)Thymus and spleen weight,spleen and bone marrowcellularity, spleen CD4+/CD5+,CD8-/CD5+and Kappa+lymphocytes0, 96, 639, or1278 mg/m3, 6hr/day, 5days/wk for 2or 4 weeksReduced spleen B-CD4+/CD5+andCD5+ T-lymphocytes at 1278 mg/m3,no effects observed at 639 mg/m3LOAEL: 1278NOAEL: 639Robinson etal., 1997ND = 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

58Early general toxicity studies reported leukopenia in dogs and fatal anem ia in m i ce exposed to 600 ppm (1917 m g /m 3 ) benzene for 12–15 days (Hough and Freem an, 1944); changes in bone m a rrow histopathology or leukopenia in rats, guinea pigs, and rabbits exposed to 80–85
ppm (256–272 m g /m 3 ) benzene for 23–187 exposures (W olf et al., 1956); and leukopenia in rats exposed to 61 ppm (195 m g /m 3 ) benzene for 2–4 weeks or to 44 ppm (141 m g /m 3 ) for 5–8 weeks (Deichm a nn et al., 1963).

Male CD-1 m i ce (11–12/group) were exposed for 6 hours/day, 5 days/week to concentrations of 0 or 10 ppm (0 or 32 m g /m 3 ) benzene for 10 weeks or to 0 or 300 ppm (0 or 958 m g /m 3 ) for 26 weeks (Green et al., 1981a,b). On the day of the last exposure, sam p les (pooled from groups of 3–4 m i ce) were obtained from the peripheral blood, bone m a rrow, and spleen to evaluate hem a tologic and hem a topoietic cells. In m i ce exposed to 10 ppm (32 m g /m 3 ), no adverse effects were observed with respect to m o rtality, body weight, or cells in the peripheral blood or bone m a rrow. Spleen weight, total nucleated cells per spleen, and nucleated RBCs per spleen were significantly increased ( p <0.05) in m i ce exposed to 10 ppm (32 m g /m 3 ). Mice exposed to 300 ppm (958 m g /m 3 ) had the following significant ( p <0.05) changes: increased m o rtality rate; decreased num bers of lym phocytes and RBCs in peripheral blood; decreased granulocyte/m acrophage progenitor cells in bone m a rrow; decreased spleen weight and num bers of lym phocytes; m u ltipotential hem a topoietic stem cells and com m itted granulocyte/m acrophage progenitor cells in the spleen; and increased incidence of atypical cell

m o rphology in the peripheral blood, bone m a rrow, and spleen. These studies identify a LOAEL of 10 ppm (32 m g /m 3 ) for slight hem a topoietic effects in m i ce exposed to benzene for 10 weeks.

In a subchronic inhalation toxicity study, W a rd et al. (1985) evaluated the peripheral blood and bone m a rrow of benzene-exposed CD-1 m i ce and Sprague-Dawley rats. Groups of 50

m a le and 50 fem a le rats and 150 m a le and 150 fem a le m i ce inhaled benzene concentrations of 0, 1, 10, 30, or 300 ppm (0, 3.2, 32, 96, or 958 m g /m 3 ) 6 hours/day, 5 days/week for up to 13 weeks. Serial sacrifice of 30 m i ce and 10 rats/group took place after 7, 14, 28, 56, and 91 days of exposure. The param e ters of toxicity evaluated included behavior, body weight, organ weights, clinical pathology, gross pathology, and histopathology. Cham b er analyses of benzene concentrations calibrated daily by gas chrom atography and m onitored continuously by infrared analyzer were within 10% of the target concentrations.

Clinical observations and body weight data revealed no signs of exposure-related toxicity. Hem a tologic effects were not observed in either species at 1, 10, or 30 ppm (3.2, 32, or 96 m g /m 3 ). The m i ce exposed to 300 ppm (958 m g /m 3 ) for 91 days, however, exhibited decreased HCT and reductions in total Hgb concentration, RBC count, W BC count, platelet count, m y eloid/erythroid ratios, and percentage of lym phocytes; the RBCs of the exposed m i ce also displayed increases in MCV, m ean cell Hgb, and glycerol lysis tim e and in the incidence and severity of red cell m o rphological changes. The values for RBC, m o tor conduction velocity,

m ean corpuscular Hgb, and glycerol lysis tim e were statistically significant ( p <0.05) for both

m a les and fem a les, whereas the rem a inder of the values were statistically significant for m ales only. Many of the effects occurred (and were statistically significant) by 14 days of exposure and m o st persisted throughout exposure, but the da ta provided indicated that the effects did not increase in severity with duration of exposure. Microscopic exam ination of the highconcentration group revealed changes in the thym us, bone m a rrow, lym ph nodes, spleen, ovaries,

59and testes. The changes in the m a les occurred m o re often and with greater severity than those in the fem a les. The m o st com m on com pound-related histopathologic findings included m y eloid hypoplasia of the bone m a rrow, lym phoid depletion of the periarteriolar sheath in the spleen and

m e senteric lym ph node, increased extram edullary hem a topoiesis in the spleen, and plasm a cell infiltration of the m a ndibular lym ph node. The incidence and/or severity of thym ic atrophy and

m y eloid hypoplasia appeared to increase with duration of exposure. The rats, not as severely affected as the m i ce, exhibited significantly ( p <0.05) decreased W BC counts (m ales on day 14 and fem a les on days 14–91) and slightly decreased fem o ral m a rrow cellularity at 300 ppm (958
mg / m 3 ).

The investigators concluded that the hem a tologic effects observed in this study are related to concentration and are sim ilar to those reported by other investigators, but that they are statistically significant only at 300 ppm (958 m g /m 3 ). In addition, the m ouse is m o re sensitive than the rat to the effects of benzene, and m a le m i ce appear to be m o re sensitive than fem a le

m i ce. The experim e nt provided a LOAEL of 300 ppm (958 m g /m 3 ) and a NOAEL of 30 ppm (96 m g /m 3 ) for peripheral blood abnorm alities in both rats and m i ce. The large num bers of anim als used in the study and the evaluation of several param e ters of hem a totoxicity strengthen the conclusions of the investigators.

4.2.1.4. Inhalation Exposure—Chronic Studies

To further characterize the hem a totoxicity of benzene, Snyder et al. (1978, 1980, 1982, 1984, 1988) conducted a series of studies using either 100 or 300 ppm (319 or 958 m g /m 3 ) benzene. These studies addressed m a inly the influence of benzene concentration and anim al species on toxic responses. The 300 ppm (958 m g /m 3 ) concentration was selected because it had produced severe hem a totoxicity in AKR m i ce exposed for life, but this concentration had produced only lym phocytopenia in Sprague-Dawley rats exposed under the sam e conditions. Also of interest was whether benzene would affect the incidence of lym phom a in the different strains of rodent. The AKR strain carries a virus that produced a high incidence of spontaneous lym phom a, killing the m i ce within 1 year, and the C57BL strain carries a virus that yields a high incidence of lym phom a following exposure to radiation, carcinogens, and im m unosuppressive agents (Snyder et al., 1980).

Exposures were sim ilar in all the studies. Essentially, test anim als and control anim als were placed in identical stainless steel inhalation cham bers and exposed to benzene or filtered air, respectively, 6 hours/day, 5 days/week for life. Benzene concentrations, m onitored every 30

m i nutes by ultraviolet absorbance at 255 nm , were read from a calibration curve. Tail bleeding of test and control anim als took place at least every 3–5 weeks during exposure and at about the

sam e tim e of day for analysis of the following endpoints: W BC 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 m a rrow. Other tissues taken for

m i croscopic exam ination included lung, liver, spleen, and kidney.

In the first study of the series, Snyder et al. (1978) exposed AKR/J m i ce (60 test and 60 control) and Sprague-Dawley rats (45 test and 27 control) to 300 ppm (958 m g /m 3 ) benzene. The rats had m ild body weight depression that started at 30 weeks and persisted throughout life.

60The only significant hem a tologic findings in the rats (± 2 standard error [SE]) were lym phocytopenia and a trend to anem ia. Lym phocyte counts for the exposed anim als were approxim a tely 75% of norm al at 2 weeks and 53% of norm al at 20 weeks. Because of increased

m o rtality, blood analyses were discontinued after 1 year of exposure. The investigators also noted fatty changes in 77% of the bone m a rrow sam p les taken from the benzene-exposed anim als com p ared with 42% of the controls. The last m ouse in the study died after 28 weeks of exposure, probably of aplastic anem ia (m ean survival tim e, 11 weeks for test anim als, 39 weeks for control anim als). The treated anim als suffere d severe weight loss (59% weight change at 16 weeks), m a rked lym phocytopenia (e.g., lym phocyte values that dropped to < 1000/m m 3 were ~3% of control levels after 36 days of exposure), bone m a rrow hypoplasia (81% of exposed m i ce 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 m o rtality. Neither species exhibited any signs of leukem i a or preleukem i a (the absence of overt signs of leukem i a, characterized by bone m a rrow dysfunction, usually low

num bers of certain cell types). The investigators suggested that this m i ght be related to early

m o rtality from aplastic anem ia in the m i ce and a lack of opportunity for recovery of the bone

m a rrow (a possible factor in leukem ogenesis) in the rats.

The investigators concluded from this study that inhalation of 300 ppm (958 m g /m 3 ) benzene causes significantly decreased survival, severe lym phocytopenia, and anem ia

accom p anied by reticulocytosis and granulosis in AKR/J m i ce and lym phocytopenia, m ild

anem ia, and m oderately decreased survival in Sprague-Dawley rats. They further concluded that AKR/J m i ce are m o re susceptible to benzene-induced hem a totoxicity than are Sprague-Dawley
rats.

The severity of hem a tologic effects induced by 300 ppm (958 m g /m 3 ) benzene in AKR/J

m i ce prom pted Snyder et al. (1980) to conduct a sim ilar study with 50 AKR/J m i ce exposed to a lower concentration (100 ppm [319 m g /m 3 ]) of benzene. The investigators also exposed 40 C57BL/6 m i ce to 300 ppm (958 m g /m 3 ) benzene to com p are the susceptibility of a different strain of m i ce with that of rats. Hem a tologic param e ters were m e asured in both strains at regular intervals during exposure.

No statistically significant differences were seen between the AKR/J test and control

m i ce in m e dian survival (39 and 41 weeks, respectively), rate of weight gain, or in type and occurrence of lym phom a. Effects that did occur were not as severe as those observed at 300
ppm (958 m g /m 3 ). But even at 100 ppm (319 m g /m 3 ), the m i ce exhibited statistically significantly reduced lym phocyte 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 m onitoring periods. Neutrophils were elevated (± 2 SE) in 3/19 m onitoring periods. Ten (20%) of the treated m i ce, but only one control, had bone m a rrow hypoplasia ( p <0.05).

C57BL/6 m i ce exposed to 300 ppm (958 m g /m 3 ) benzene had a m e dian survival tim e of 41 weeks, com p ared with 75 weeks for the controls. Starting 1 week after the beginning of
exposure, m i ce exposed to benzene exhibited a statistically significant increase in incidence of lym phocytopenia at 29 of 30 m onitoring periods (W BC counts were about 3000/m m 3 or ~15% of control values), and of anem ia in 30 of 30 m onitoring periods (RBC counts were ~80% of

61control values). After 17 weeks, the incidence of neutrophilia was significantly increased in 15 of the 22 rem a ining m onitoring periods. Throughout exposure, the peripheral RBCs exhibited anisocytosis and poikilocytosis. A neutrophilic shift to the left, characterized by the appearance of m e tam y elocytes, indicated im m a turity. At the end of the first year, m y elocytes,
prom yelocytes, and giant platelets began to appear. Blood analyses were discontinued after 61 weeks because of decreased survival. Exam ination of the bone m a rrow revealed hyperplasia,

m a inly of granulopoietic elem ents, in 13 (33%) of the exposed m i ce; hyperplasia was not present in the controls. There was a significant increase in the incidence of hem a topoietic neoplasm s , including six cases (15%) of thym ic lym phom a. Two control m i ce died with nonthym ic lym phom a.

The investigators concluded from this study that the two strains of m i ce are m o re sensitive than the Sprague-Dawley rat to the hem a tologic effects of benzene and that the AKR/J

m ouse is m o re sensitive than the C57BL/6 m ouse. The response of the C57BL/6 m ouse producing a proliferative effect on the m y eloid cell line is significant, because this is the cell line that undergoes leukem i c transform ation in hum ans exposed to benzene.

In the next study, Snyder et al. (1982) exam ined the effects of 300 ppm (958 m g /m 3 ) benzene on CD-1 m i ce, a strain not known to harbor any endogenous lym phom a virus. The effects were sim ilar to those observed in the AKR/J and C57BL/6 strains and included decreased

m e dian survival tim e (25.5 weeks for benzene-exposed anim als, 52.7 weeks for controls), statistically significantly depressed peripheral RBC and lym phocyte 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 (rem n ants of nuclear chrom atin seen in erythrocytes in certain anem ic 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 im m a ture m y eloid cells at 217 days.

Tum o r incidence was not significant. One benzene-exposed anim al developed acute

m y eloblastic leukem i a, one had chronic m y elogenous leukem i a, one had a benign lung adenom a, and two had m a lignant lym phom a with thym ic involvem e nt. Of the 35 benzene-exposed anim als that did not have neoplasia, 9 had bone m a rrow hyperplasia, 11 had bone m a rrow hypoplasia, 6 had splenic hem o siderin pigm ents (indicating hem o lysis or ineffective erythropoiesis), and 19 had splenic hyperplasia. The investigators noted that “exposed m i ce dying with bone m a rrow hypoplasia survived, on the average, 75 fewer days than m i ce dying with bone m a rrow hyperplasia. This m a y indicate that bone m a rrow hypoplasia is an early response to exposure followed by bone m a rrow hyperplasia. It m a y also indicate two different responses to the exposure.”

Snyder et al. (1988) exam ined the influence of m odifications in inhalation protocol on benzene-induced hem a totoxicity and tum o r developm ent in C57BL/6 and CD-1 m i ce. Exposure was by two protocols, one representing interm ittent occupational exposures (1 week of exposure to 300 ppm [958 m g /m 3 ] benzene followed by 2 weeks of nonexposure alternately for life), and the other representing intense but short-term exposures (1200 ppm [3834 m g /m 3 ] benzene for 10 weeks); all exposures were 6 hours/day, 5 days/week. The long-term interm ittent exposures produced earlier m o rtality in both strains of m i ce than did the short-term
, 10-week exposures. Both types of exposure produced severe lym phocytopenia and m oderate anem ia in both strains

62of m i ce, but peripheral blood values in the 1200 ppm (3834 m g /m 3 )/10-week exposure groups returned to norm al after term ination of exposure. The 300 ppm (958 m g /m 3 ) exposure induced lym phocytopenia and anem ia throughout the course of exposures. Tum o r incidences were significantly increased ( p <0.05–0.001) with both protocols in CD-1 m i ce and with the interm ittent, long-term exposure in the C57BL/6 strain. A puzzling finding was that neither leukem i a nor lym phom a occurred in either strain by either protocol. The investigators concluded that benzene was cytotoxic by either protocol, especially to circulating lym phocytes (but recovery took place when exposure to the high concentration ceased) and that both protocols did induce tum o rs. The Snyder et al. (1978, 1980, 1982, 1984, 1988) studies were designed to answer specific questions and used only one concentration at a tim e (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 abnorm alities reflect a disruption of hem a topoiesis in the bone m a rrow; thus m a ny experim e ntal studies have focused on this process, analyzing the total cellularity (the total num ber of bone m a rrow cells obtained from a fem u r), as well as the individual cellular com ponents of the bone m a rrow. The pluripotent hem a topoietic
stem cells are capable of self-renewal and can differentiate along at least three lineages: erythrocytic, granulocytic/m acrophagic, and throm bocytic (Snyder, 1987). These cells are quantitated by the spleen colony assay in which the colonies form ed by the cells in the spleens of lethally irradiated host anim als are counted; these are called colony-form ing units–spleen (CFUS). The specific progenitor cells, in contrast to the pluripotential stem cells, are com m itted to differentiate, and they also have som e capacity for self-renewal (cell division). These cells are

enum erated by counting the colonies form ed in cell cultures grown in the presence of specific growth-factor stim uli (Snyder, 1987). The two types of erythroid progenitor cells are the lessdifferentiated burst-form ing units–erythroid (BFU-E), which responds to a factor called “burstprom oting activity” or to IL-2, and the m o re differentiated erythroid progenitor, the colonyform ing units–erythroid (CFU-E), which responds to erythropoietin. The granulocyte/m acrophage colony-form ing units–culture (GM-CFU-C) respond to conditioned

m e dia from hum an leukem i a cells or from organs of anim als treated with plant lectins. The Tand B-lym phocytes arise from the pluripotential hem a topoietic stem cells in the bone m a rrow. The cells progress to the recognizable blast cells and their progeny and further m a ture to the functional form s norm ally observed in peripheral blood (Snyder, 1987).

Seidel et al. (1989) exam ined the effects of benzene on peripheral blood cell counts and the hem a topoietic stem cell com p artm ents of the bone m a rrow. Fem a le BDF1 m i ce were exposed to 100, 300, or 900 ppm (319, 958, or 2875 m g /m 3 ) benzene 6 hours/day, 5 days/week for 16 weeks. Benzene concentrations were m onitored twice a day. Control m i ce, placed in cham bers, were exposed to air. The evaluation of hem a topoietic effects included W B C (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 lym phocyte count after both 4 and 8 weeks of
exposure. Lym phocyte counts of the anim als exposed to 300 and 900 ppm (958 and 2875

63mg / m 3 ) returned to control values 1–2 weeks after cessation of exposure, whereas the lym phocytes in the 100 ppm (319 m g /m 3 ) group rem a ined low during that tim e. The anim als exposed to 300 and 900 ppm (958 and 2875 m g /m 3 ) developed a slight anem ia that did not worsen with longer exposure. Three-fourths of the anim als with elevated CFU-E in the bone

m a rrow also showed reticulocytosis at 13 weeks. The num ber of granulocytes was apparently not affected. The investigators noted that control values for the 300 ppm (958 m g /m 3 ) groups were approxim a tely twice those for the 100 and 900 ppm (319 and 2875 m g /m 3 ) groups, possibly because the 300 ppm (958 m g /m 3 ) experim e nt had been perform ed 1 year earlier.

Assays perform ed after 4 and 8 weeks of exposure dem onstrated decreases in stem cell

num bers that were dose related, particularly at 8 weeks. At 4 weeks, the CFU-E count was depressed m o re than other stem cell counts. At 100 ppm (319 m g /m 3 ), the CFU-E were depressed to 48% of control, whereas the BFU-E, CFU-S, and CFU-C (colony-form ing unit–culture) were depressed to 88%, 92%, and 94% of control, respectively. At 8 weeks, the values for CFU-E in anim als exposed to 100, 300, and 900 ppm (319, 958, and 2875 m g /m 3 ) were 99%, 73%, and 35% of controls, respectively. Sim ilar patterns of depression were observed for BFU-E and CFU-C num bers of anim als exposed to benzene for 8 weeks, and CFUS num bers were depressed to 65% of control at 300 ppm (958 m g /m 3 ), the only dose tested after 8 weeks of exposure. Recovery studies dem onstrated a gradual regeneration of progenitor cell

num bers, and anim als exposed to 300 ppm (958 m g /m 3 ) for 16 weeks had com p letely recovered by 73–185 days postexposure. The investigators concluded that the erythroid system is m o re sensitive than the m y eloid cell system to the effects of benzene, with the CFU-E showing m o re sensitivity than other com ponents of the erythroid system under the conditions of this study.

Two studies were designed to exam ine the effects of low concentrations of benzene on the erythroid and m y eloid progenitor cells (Baarson et al., 1984; Baarson and Snyder, 1991). Baarson et al. (1984) exam ined the effects of subchronic exposure to a low concentration of benzene on the erythroid progenitor cells in the bone m a rrow and spleen of m i ce. Male C57BL/6J m i ce inhaled 10 ppm (32 m g /m 3 ) benzene vapor 6 hours/day, 5 days/week for as long as 178 days. Peripheral blood counts and bone m a rrow and spleen BFU-E and CFU-E assays were perform ed on days 32, 66, and 178 of exposure. The num bers of circulating RBCs in the benzene-treated groups were significantly decreased ( p <0.05) at both 66 and 178 days. Lym phocyte values were decreased ( p <0.05) at all three tim e points. Benzene did not affect levels of circulating neutrophils, bone m a rrow cellularity, and num bers of nucleated red cells in the bone m a rrow. The num bers of bone m a rrow 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 tim e points, was apparently exponential with a half-life of about 2 m onths. Bone m a rrow BFU-E colonies were significantly depressed in benzene-exposed anim als (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 num bers (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 m i ce 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 tim e points. The spleen is a site of
extram edullary erythropoiesis in the rodent, particularly under stressful conditions. The reduction in the num bers of nucleated red cells and CFU-E in the spleen following exposure to benzene indicated that the com p ensatory m echanism of the spleen was also affected. The

64investigators concluded that low concentrations of benzene are hem a totoxic and that further studies to exam ine this effect would be warranted.

In the later study, Baarson and Snyder (1991) reported that the ingestion of ethanol in

com b ination with exposure to benzene m a y enhance benzene-induced toxicity of the erythroid progenitor cells in C57BL/6J m i ce. These studies identify a LOAEL of 10 ppm (32 m g /m 3 ) for depressed hem a topoiesis in m i ce.

Vacha and coworkers (1990) exposed fem a le C57BL/6×DBA/2F1 hybrid m i ce to 0 or 300 ppm (0 or 958 m g /m 3 ) benzene 6 hours/day, 5 days/week for 6–7 weeks. Indices of hem a topoiesis were m easured in peripheral blood (RBC and W BC count, Hgb, HCT, reticulocyte, and leukocyte count) in addition to 59 Fe accum u lation in the erythropoiesis organs (spleen and bone m a rrow) and in the peripheral RBCs. The distribution of developm ental classes of erythroblasts was also determ ined. This study found that anim als becam e anem ic after 6–7 weeks of benzene exposure. The num ber of erythroblasts in the bone m a rrow was not different; however, exposure to benzene shifted the population to a less m a ture class of cells. The num ber of colonies derived from BFU-E and CFU-E were decreased to 70 and 34% of controls, respectively. A LOAEL of 300 ppm (958 m g /m 3 ) was established for hem a totoxic effects.

Toft et al. (1982) evaluated the adverse effects of occupationally relevant levels of benzene on the bone m a rrow of m i ce. Male NMRI m i ce inhaled concentrations of benzene that ranged from 1 to 200 ppm (3.2 to 639 m g /m 3 ). Exposures were either continuous (24 hours/day for 4 to 10 days) or interm ittent (8 hours/day, 5 days/week for 2 weeks) using five m i ce per group. Endpoints included the num ber of nucleated cells per tibia, the number of colonyform ing granulopoietic stem cells (CFU-C/tibia), and frequency of m i cronuclei in polychrom atic erythrocytes (PCE). Mice exposed continuously to 

###  21 ppm (67 m g /m 3 ) benzene exhibited significant and concentration-dependent alterations in all three param e ters. The values after 4 days of exposure to 21 ppm (67 m g /m 3 ) of benzene (estim ated from a graph) were cells/tibia, ~24% of control; CFU-C/tibia, ~32% of control; m i cronuclei/500 PCE, 6 (control value, 0.41/500).

Furtherm ore, interm ittent inhalation exposure to 1.0, 10.5, 21, 50, 95, and 107 ppm (3.2, 33, 67, 160, 303, and 342 m g /m 3 ) for 2 weeks produced concentration-dependent decreases in CFU-C/tibia and increases in the frequency of m i cronuclei, significant at 

###  21 ppm (67 m g /m 3 ), and concentration-dependent increases in cellularity, significant at 

###  50 ppm (160 m g /m 3 ). Anim als exposed interm ittently to 95 or 201 ppm (303 or 642 m g /m 3 ) benzene 2–8 hours/day, 5 days/week for 2 weeks had significant decreases in cellularity and CFU-C/tibia (95 ppm [303
mg / m 3 ] for 6 and 8 hours/day) and increases in the frequency of m i cronuclei (95 ppm [303
mg / m 3 ] for 4 hours). At 201 ppm (642 m g /m 3 ), all exposure protocols, except for the 2 hours/day exposure, produced significantly adverse effects on all param e ters. Statistical significance was determ ined by the Student’s t test for cellularity and CFU-C and by the W ilcoxW h itney m e thod for m i cronuclei, which was set at p 

###  0.05. 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 m o st exposures and that CFU-C are m o re sensitive to prolonged

65exposure to low levels of benzene than are the m a jority of bone m a rrow 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 m a rrow but that the injury could be offset by a com p ensatory increase in the proliferation rate of the bone m a rrow. The investigators further concluded that interm ittent exposure is less effective than continuous exposure, that the induction of m i cronuclei is im portant because som a tic m u tations precede m o st chem ically induced cancers, and that the cytotoxicity observed in the study could lead to aplastic

anem ia.

Because of the relationship between exposure to benzene and the developm ent of

m y eloblastic leukem i a, Dem p ster and Snyder (1990) evaluated the possibility that exposure provides a growth advantage for granulocytic progenitor cells over erythroid progenitor cells. Basically, DBA/2 m a le m i ce inhaled benzene concentrations of 0, 10, 30, or 100 ppm (0, 32, 96, or 319 m g /m 3 ) 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 determ ine the effects of benzene on stem cells during increased erythropoiesis, subgroups of five exposed m i ce received injections of hem o lytic doses of phenylhydrazine during or after exposure. Controls were exposed to air and/or injected with saline. Brief descriptions of treatm e nt 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

m a rrow BFU-E and CFU-E, but GM-CFU-C were not affected (except for a spurious increase at 30 ppm [96 m g /m 3 ]). Splenic CFU-E and BFU-E were increased at 10 ppm (32 m g /m 3 ) and CFU-E were depressed at 100 ppm (319 m g /m 3 ). Splenic granulocytic cells were unaffected. Five days after exposure, BFU-E and CFU-F were the sam e as in controls, but GM-CFU-C were decreased at 10 and 100 ppm (32 and 319 m g /m 3 ); splenic CFU-E exhibited a concentrationrelated increase, significant at 30 and 100 ppm (96 and 319 m g /m 3 ), whereas the num bers of the other two progenitor cell types were norm al.

Benzene + phenylhydrazine. One day after exposure, bone m a rrow BFU-E were depressed at 30 and 100 ppm (96 and 319 m g /m 3 ), and CFU-E were depressed only at 30 ppm (96 m g /m 3 ); there was no effect on the num ber 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 m g /m 3 ). Five days after exposure, bone m a rrow CFU-E were elevated in m i ce exposed to 10 ppm (32 m g /m 3 ), and GM-CFU-C exhibited a concentration-dependent depression; bone

m a rrow BFU-E counts were sim ilar to control counts. Splenic GM-CFU-C were elevated at all concentrations, CFU-E were elevated at 10 and 30 ppm (32 and 96 m g /m 3 ) and depressed at 100
ppm (319 m g /m 3 ), and BFU-E were elevated only at 10 ppm (32 m g /m 3 ).

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 m a rrow and the spleen of exposed m i ce. This observed shift toward granulopoiesis occurred even in m i ce treated with the erythropoiesis stim ulus phenylhydrazine, but the effects were short-lived. The bone m a rrow erythroid progenitor cells had recovered from their depression 5 days after exposure, and at that tim e the granulocytic progenitor cells had becom e depressed. The dose-dependent increase in splenic CFU-E 5 days after exposure probably reflects the spleen’s attem p t to repopulate the erythron.

66In a series of studies, Luke et al. (1988a, b) and Tice et al. (1989) exam ined the effect of sex, route, schedule, and duration of exposure on benzene-induced cytotoxicity and genotoxicity of the bone m a rrow in m i ce. The effects on percentage of PCE in peripheral blood, PCV, and bone m a rrow cellularity (all considered by the investigators to be m easures of bone m a rrow cytotoxicity) are reviewed below.

Groups of 6–10 m a le and fem a le DBA/2 m i ce, m a le B6C3F1 m i ce, and m a le C57BL/6J

m i ce inhaled either 300 ppm (958 m g /m 3 ) benzene or am bient air 6 hours/day for 13 weeks. Exposures took place on either 3 consecutive days/week or 5 consecutive days/week. For

com p arison with the oral route of exposure, one group of B6C3F1 m a les received doses of 400

m g /kg by gavage on 5 consecutive days/week for 14 weeks. The oral dose of 400 m g /kg was estim ated to exceed the total am ount of benzene absorbed by a m ouse during a 6-hour exposure to 300 ppm (958 m g /m 3 ). Peripheral blood sm ears, prepared weekly, were used to exam ine PCE. PCVs and bone m a rrow cellularity were determ ined at the end of the study. The data for the three param e ters were statistically analyzed using tem poral averages and then evaluated for differences related to sex, strain, regim e n, and route of exposure using a two-way Brown Forsythe analysis of variance (ANOVA). Group m ean data were com p ared using Student’s t test.

Benzene initially induced a significant depression in the num ber of PCE in the peripheral blood of all three strains exposed by either route and both inhalation regim e ns. However, the extent and duration of the depression varied by sex, strain, and exposure regim e n. Fem a le DBA/2 m i ce exhibited the sm allest initial decrease in PCE, and values returned to control levels by week 3 of exposure. The extent of depression was not dependent on exposure regim e n. Am ong the m a les of the different strains, the DBA/2 m ouse was the m o st 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 m i ce) and exposure regim e n. Oral treatm e nt produced an initial suppression in PCE levels in the m a le B6C3F1 m i ce that persisted for only 2 weeks and was not revealed in the
tem poral average.

The study also dem onstrated a significant depression in the PCV and bone m a rrow cellularity (determ ined at the end of the study) of the exposed m i ce. W ith regard to the PCV values, the depression was independent of the inhalation regim e n in B6C3F1 and C57BL/6 m i ce. After correcting for differences in PCV levels in controls, the data showed that inhalationproduced 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 m i ce the extent of depression was sex dependent (m ale > fem a le; p <0.025). W ith regard to bone m a rrow cellularity, the depression observed after inhalation exposure depended on sex (m ale > fem a le;
p <0.0002) and strain (DBA/2 < B6C3F1 

###  C57BL/6; p <0.0012). Am ong the m a le m i ce, the extent of depression was dependent on route (inhalation > oral; p <0.012) and was not dependent on regim e n.

These studies used rather sm all num bers of anim als but otherwise appear to have been thoroughly conducted and evaluated; however, the inform ation provided is for exposure to benzene that is significantly above the m o st recent occupational lim its. The investigators

67suggest that further studies are needed to evaluate the influences on benzene toxicity relative to regim e n and route of exposure at lower doses.

In a series of experim e nts, Cronkite et al. (1985, 1989) exam ined the hem a totoxicity of inhaled benzene in m i ce. The study exam ined the effects on and recovery of peripheral blood, bone m a rrow, and progenitor cells, as well as the developm ent of neoplasia and the influence of exposure regim e n on toxicity. In the first study, m a le and fem a le C57BL/6 BNL m i ce were exposed for 6 hours/day, 5 days/week to benzene concentrations of 10, 25, 100, or 400 ppm (32, 80, 319, or 1278 m g /m 3 ) for 2 weeks or to 300 ppm (958 m g /m 3 ) for durations of exposure ranging from 2 to 16 weeks. The endpoints of toxicity included W BC, RBC (Coulter), and differential counts; HCT; bone m a rrow cellularity; CFU-S in the bone m a rrow and spleen; and
tum o r developm ent. Assays were perform ed 18–22 hours after term ination of exposure or after various periods of recovery up to 16 weeks.

Inhalation of 25, 100, or 400 ppm (80, 319, or 1278 m g /m 3 ) benzene for 2 weeks did not affect granulocyte counts but did result in a concentration-related depression in the num ber of lym phocytes in the peripheral blood; there was no effect on lym phocytes at 10 ppm (32 m g /m 3 ). HCT values were decreased at 100 and 400 ppm (319 and 1278 m g /m 3 ); RBC counts were also decreased, but no data were given. At the end of exposure, lym phocyte depression ranged from about 20 to 65% of control for the three highest exposure groups. For m i ce exposed to 300 ppm for 2 weeks, lym phocyte levels were sim ilar to control levels by 4 weeks after the end of exposure, and recovery of the rem a ining exposure duration groups was com p lete by 8 weeks after exposure. This experim e nt provided an estim ated LOAEL of 25 ppm (80 m g /m 3 ) and a NOAEL of 10 ppm (32 m g/m 3 ) for lym phocyte depression in m i ce exposed to benzene for 2 weeks (Cronkite et al., 1985). These results differed from those of W a rd et al. (1985), which suggested a LOAEL of 300 ppm (958 m g /m 3 ) for all hem a tologic param e ters tested in both m i ce and rats. The m a in difference in the two studies was the use of different strains of m ouse.

Cronkite et al. (1985) also determ ined bone m a rrow cellularity, the percentage of bone

m a rrow stem cells synthesizing DNA, num bers of pluripotent stem cells in bone m a rrow, and CFU-S. Cellularity was determ ined in anim als inhaling 0, 10, 25, or 100 ppm (0, 32, 128, or 319
mg / m 3 ) for 2 weeks (the 400 ppm [1278 m g /m 3 ] concentration was not tested as it was for the other param e ters of the study). Benzene levels of 100 ppm (319 m g /m 3 ) significantly depressed bone m a rrow cellularity and stem cell CFU-S num bers ( p <0.003 and p <0.001, respectively) in exposed m i ce. In contrast, 10 and 25 ppm (32 and 128 m g /m 3 ) had no effect on these
param e ters. A single assay dem onstrated a depression in the fraction of stem cells in DNA synthesis in m i ce exposed to 10 ppm (32 m g /m 3 ). There was no effect at 25 ppm (128 m g /m 3 ), and slight increases in the fraction in DNA synthesis at 100 and 400 ppm (319 and 1278 m g /m 3 ) ( p <0.17 and p <0.08, respectively). These results for DNA synthesis are of uncertain relevance.

In another experim e nt, the recovery of bone m a rrow CFU-S population was tracked after exposure of the anim als to 300 ppm (958 m g /m 3 ) for 2, 4, 8, or 16 weeks. In anim als exposed for 2 and 4 weeks, bone m a rrow CFU-S num bers 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 anim als 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 anim als exposed for 16 weeks, CFU-S values were depressed to 27% of

68control, and although recovery was occurring by week 2 after the last exposure, it was still not

com p lete by 16 weeks after term ination of exposure. The investigators concluded that exposure to 300 ppm (958 m g /m 3 ) benzene for 16 weeks significantly depressed the hem opoietic stem cells and that recovery was incom p lete 16 weeks after term ination of exposure.

In reviewing the effects of benzene on the CFU-S com p artm ent, Snyder (1987) pointed out that under norm al circum stances the CFU-S cells are nondividing and in that state are probably resistant to benzene toxicity; however, in response to benzene-induced injury of m o re differentiated cell types, the CFU-S m a y be drawn into the cell cycle and thereby becom e susceptible to benzene.

Cronkite et al. (1985) also noted that anim als exposed to 300 ppm (958 m g /m 3 ) benzene for 16 weeks began to die from thym ic and nonthym ic lym phom ata and solid tum o rs, 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

m acrophage progenitors (GM-CFU-C), (2) com p ared the effects of two different exposure regim e ns on peripheral blood cell counts and progenitor cell num bers, (3) evaluated m o rtality and neoplasia, and (4) tested the functional capacity of the stem cells using a “rescue assay.” Hale-Stoner m i ce, aged 12–14 weeks, were used in the progenitor cell assays. Bone m a rrow cell suspensions harvested from m i ce exposed to 400 ppm (1278 m g /m 3 ) benzene 6 hours/day, 5 days/week for 9.5 weeks and from their corresponding controls were counted and incubated to form plasm a 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-m itogen conditioned m e dium

. The clots were scored for erythrocytic colonies (CFU-E) or

m e gakaryocytes (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 num bers did not change significantly after 4 days of exposures, but the num bers had decreased significantly (to approxim a tely 45–70% of control values) by assay days 29, 48, and 65. Twelve days after exposure ended, the cell num bers 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 tim e of this assay, the peripheral RBC counts were decreased to about 50–75% of control values.

Granulocyte/m acrophage 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 approxim a tely 40% of control values after 30 days of exposure, increased to about 75% by term ination of exposure at day 65, decreased to approxim a tely 60% by 72 days, and then recovered to alm o st 80% by 80 days. The num bers of granulocytes in the blood also fluctuated during this tim e.

Influence of exposure protocols. CBA/Ca BNL m a le m i ce were exposed to either 316
ppm (1010 m g /m 3 ) benzene adm i nistered 6 hours/day, 5 days/week for 19 exposures or to 3000
ppm (9584 m g /m 3 ) benzene adm i nistered 6 hours/day for 2 successive days (Cronkite et al., 1989). After 19 exposures, 316 ppm (1010 m g /m 3 ) benzene had reduced lym phocyte counts from 7500/ 

###  L to 300/ 

###  L, and 3000 ppm (9584 m g /m 3 ) 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

69the two regim e ns still persisted at 214 days after exposure. Neutrophil counts of the group exposed to 316 ppm (1010 m g /m 3 ) were also depressed m o re than those exposed to 3000 ppm (9584 m g /m 3 ). By 214 days after exposure, the neutrophil counts in the group treated with 3000
ppm (9584 m g /m 3 ) had recovered, whereas those in the group treated with 300 ppm (958 m g /m 3 ) had not. All leukocyte cell types were decreased in num ber as com p ared with those in sham

exposed m i ce on day 1, and all absolute counts except for the large unstained cells (m ostly large lym phocytes) were depressed on 32, 67, and 214 days after the last exposure. These data were not given for the anim als exposed to 3000 ppm (9584 m g /m 3 ).

On day 1 after exposure, both treatm e nts had induced a reduction in bone m a rrow cellularity; this was significantly greater in the anim als exposed to 316 ppm (1010 m g /m 3 ) for 19 exposures (depression to ~33% of control value) than in those exposed to 3000 ppm (9584
mg / m 3 ) 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 sam e response pattern except that by day 214 after exposure recovery was incom p lete for the group exposed to 316 ppm (958 m g /m 3 ).

In another experim e nt, exposure to 300 ppm (958 m g /m 3 ) benzene for 80 weeks produced a severe depression in CFU-S (to 38% of control) that had not com p letely recovered 178 days after exposure ended. Exposure to 3000 ppm (9584 m g /m 3 ) for 8 weeks produced a m ilder depression in CFU-S (to 62% of control) that had recovered by 30 days. Decreases in bone

m a rrow cellularity were nearly the sam e for the two exposure scenarios.

Rescue assay. The injection of norm al bone m a rrow 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 m a rrow. C57BL/6BNL m a le and fem a le m i ce were irradiated with 850 rad, then injected with bone m a rrow cells from m i ce that had been exposed to 300 ppm (958 m g /m 3 ) benzene 6 hours/day, 5 days/week for 16 weeks 752 days earlier. Thirty days after the initial transplants, som e of the survivors’ m a rrow was injected into another group of irradiated m i ce for a secondary transplant, and 30 days later those recipients served as donors in the third transplant to still another group of irradiated m i ce. The recipients of bone m a rrow 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 m i ce that received cells from control m i ce was 93%, and survival of the m i ce that received cells from benzene-exposed

m i ce was 64%.

In the secondary transplant, no m i ce died by 30 days, but the m i ce that received control cells started dying at day 171 and m i ce that received cells from donors exposed to benzene started to die at day 70. In the tertiary rescue, 50% of the recipients of bone m a rrow from benzene-exposed donors were dead by day 30 and 95% were dead by day 245. Am ong the anim als that received cells from the sham -exposed m i ce, only one died by day 30 and 50% died by day 375. The last anim al died on day 594. The m echanism for the stem cell m a lfunction that led to increased m o rtality is not clear, but the investigators suggest that the observation could be associated either with genetic injury to the stem cells, exhaustion of the G 0 hem a topoietic stem cells, or a radiation effect on the strom a of the recipient.

70BDF1 m i ce were exposed to 0, 100, 300, or 900 ppm (0, 319, 958, or 2875 m g /m 3 ) benzene for up to 4 weeks (Seidel et al., 1990). The num bers of hem a topoietic progenitor cells, early and late progenitors (BFU-E, CFU-E), and granuloid progenitors (CFU-C) were determ ined. A group was generated to establish the effect of ethanol (drinking water) on these effects. This study dem onstrated that benzene decreased the num ber of CFU-E per fem u r in a concentration-dependent m a nner. This effect was evident at 300 and 900 ppm (958 and 2875
mg / m 3 ) concentration; however, the effect in the 100 ppm (319 m g /m 3 ) exposure group was uncertain, as the study focused on the effects of ethanol on benzene toxicity. The LOAEL/NOAEL was thus difficult to determ ine.

Male Sprague-Dawley rats (40/group) were exposed to vapor concentrations of 0 or 100
ppm (0 or 319 m g /m 3 ) benzene 6 hours/day, 5 days/week for life (Am e rican Petroleum Institute, 1983). Blood sam p les were obtained at 2- to 4-week intervals throughout the treatm e nt period. The erythrocyte and lym phocyte counts were depressed at nearly every sam p ling tim e 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, W ilcoxon Rank Sum Test), and the overall results were interpreted as evidence of hem a totoxicity. Significantly increased incidences of splenic hyperplasia ( p <0.005) and hem o siderin pigm ents ( p <0.001) were observed in benzene-exposed rats. The incidences of norm ally rare tum o rs in treated rats were 4/40 in liver, 2/40 in Zym b al gland, and 1/40 chronic m y elogenous leukem i as. The authors considered these tum o rs to be related to the benzene exposure. This study identifies a LOAEL of 100 ppm (319 m g /m 3 ) for hem a tologic effects in rats.

Decreases in bone m a rrow cellularity in the fem u r, HCT, and leukocytes were seen in DBA/2 m i ce (20/group) exposed to 300 ppm (958 m g /m 3 ) benzene in air 6 hours/day, 5 days/week for 2 weeks (Chertkov et al., 1992). In m o st cases no erythroid or m y eloid clonogenic cells could be recovered in bone m a rrow cultures started after the last benzene
exposure. After 2 weeks of recovery, however, body weight, HCT, bone m a rrow cellularity, and

com m itted hem a topoietic progenitor cells had recovered to near norm al values.

Neun et al. (1992, 1994) com p ared the in vivo sensitivity of Swiss-W e bster and C57B1/6J m i ce to inhaled benzene with the in vitro sensitivity of culture bone m a rrow cells to benzene m e tabolites. Mice were exposed to 300 ppm (958 m g /m 3 ) benzene 6 hours/day, 4 days/week for 2 weeks. Swiss-W e bster m i ce were m o re sensitive to benzene exposure, as indicated by m u ch greater reductions in fem o ral bone m a rrow cellularity and in the num ber of CFU-E per fem u r after in vivo exposure. Neither phenol nor m u conic acid were toxic to cultured CFU-E from either m ouse strain. CFU-E from Swiss-W e bster m i ce were m o re sensitive than CFU-E from C57B1/6J m i ce to 1,4-benzoquinone or hydroquinone. Thus, both in vitro and in vivo data indicated that Swiss-W e bster m i ce were m o re sensitive to benzene toxicity than were C57B1/6J m i ce.

Fem a le BDF1 m i ce (C57BL/6 × DBA/2F1 hybrids) were exposed to 0, 100, 300, or 900
ppm (0, 319, 958, or 2875 m g /m 3 ) benzene 6 hours/day, 5 days/week for up to 8 weeks (Plappert et al., 1994). Hem a tologic studies included peripheral blood data, T4 and T8 lym phocyte counts in the blood and spleen, and hem a topoietic stem and progenitor cell assays in the m a rrow CFUS, CFU-C, and BFU-E, CFU-E. No significant changes were observed in the peripheral blood data of m i ce exposed to 900 ppm (2875 m g /m 3 ) benzene. Som e perturbation of the reticulocyte

71num bers was observed, but values at days 3 and 5 did not differ from controls. Absolute

num bers of lym phocytes and neutrophils did not differ from controls (no data shown). Slight

anem ia was observed at 4 and 8 weeks of treatm e nt with 300 and 900 ppm (958 and 2875
mg / m 3 ) 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-form ing cell

num ber appeared at 4 and 8 weeks of exposure, with m a xim a l effect at the level of CFU-E.

Farris et al. (1997a) exposed m a le B6C3F1 m i ce to 0, 1, 10, 100, or 200 ppm (0, 3.2, 32, 319, or 639 m g /m 3 ) benzene 6 hours/day, 5 days/week for 1, 2, 4, or 8 weeks. A separate 4-week experim e nt at benzene concentrations of 0, 1, 5, and 10 ppm (0, 3.2, 16, and 32 m g /m 3 ) was also conducted. Another group of anim als was exposed to benzene for 4 weeks and then allowed to recover for up to 25 days. Bone m a rrow cell counts, high proliferative potential colony-form ing unit (CFU-HPP), percentage of CFU-HPP cells in S-phase, bone m a rrow CFU-E, bone m a rrow GM-CFU, labeling index, B and T lym phocytes, and hem a tology param e ters were determ ined. Data sets for each param e ter were individually evaluated with a univariate two-way ANOVA using contrasts to determ ine treatm e nt effect. Statistical significance was determ ined at a level of p <0.05.

There were no significant effects on hem a topoietic param e ters from exposure to 10 ppm (32 m g /m 3 ) benzene or less. Exposure of m i ce to 100 or 200 ppm (320 or 640 m g /m 3 ) reduced the num ber of total bone m a rrow cells, progenitor cells, differentiating hem a topoietic cells, and

m o st blood param e ters, with a concentration-related effect. Replication of prim itive progenitor cells in the bone m a rrow was increased during the exposure period, possibly in com p ensation for benzene cytotoxicity. Recovery after benzene exposure was rapid. Most hem a topoietic
param e ters returned to control levels within 4 days following the 100 ppm (320 m g /m 3 ) exposure and within 11 days following the 200 ppm (640 m g /m 3 ) 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 m i cronucleated erythrocytes was increased in B6C3F1 m a le m i ce exposed to 100 and 200 ppm (320 and 640 m g /m 3 ) benzene for 8 weeks. The authors suggested that the increased proliferation of prim itive progenitor cells, in concert with genetic dam a ge, provides the com ponents for producing an increased incidence of lym phom a in m i ce. The effects on hem a topoietic param e ters indicate a LOAEL of 100 ppm (320 m g /m 3 ) and a NOAEL of 10 ppm (32 m g /m 3 ).

Farris et al. (1997b) exposed m a le B6C3F1 m i ce to 0, 1, 10, 100, or 200 ppm (0, 3.2, 32, 319, or 639 m g /m 3 ) benzene 6 hours/day, 5 days/week for 1, 2, 4, or 8 weeks. A separate 4-week experim e nt at benzene concentrations of 0, 1, 5, and 10 ppm (0, 3.2, 16, and 32 m g /m 3 ) was also conducted. Spleen and thym ic lym phocyte counts were determ ined for three m i ce in each
treatm e nt. Fem o ral B lym phocyte counts were evaluated in 10 m i ce per group. Labeling index of bone m a rrow B lym phocytes was determ ined by BrdU incorporation. To m i nim i ze the potential for recovery, m i ce were sacrificed and sam p led within 2 hours of the end of the last
exposure. The spleen, thym us, and bone m a rrow data sets were individually evaluated with a univariate two-way ANOVA using contrasts to determ ine treatm e nt effect within each tim e point. Statistical significance was determ ined at a level of p <0.05.

Exposure to 100 or 200 ppm (319 or 639 m g /m 3 ) benzene induced rapid and persistent reductions in fem o ral B, splenic T and B, and thym ic T lym phocytes; total nucleated thym us

72cells count; and total nucleated spleen cells. There was a significant decrease in thym ic cell count and B lym phocytes in the spleen at week 2 in m i ce exposed to 10 ppm (32 m g /m 3 ) in the 8-week study and no effects at 1 ppm (3.2 m g /m 3 ). These param e ters, however, were

com p arable with the controls by week 4 of the 8-week exposure experim e nt, and in the separate 4-week experim e nt there were no significant decreases in these param e ters at any tim e in the

m i ce exposed to 10 ppm (32 m g /m 3 ) benzene or less. Thus, the 100 ppm (319 m g /m 3 ) level was considered a LOAEL and 10 ppm (32 m g /m 3 ) a NOAEL for these effects. The percentage of fem o ral B lym phocytes and thym ic T lym phocytes in apoptosis was increased 6- to 15-fold by 200 ppm (639 m g /m 3 ). Replication of fem o ral B lym phocytes was increased during the exposure period in the bone m a rrow, apparently in com p ensation for the lym phocyte loss induced by 100 or 200 ppm (319 or 639 m g /m 3 ) benzene exposure.

Robinson and coworkers (1997) exposed m a le Sprague-Dawley rats (16/group) to 0, 30, 200, or 400 ppm (0, 96, 639, or 1278 m g /m 3 ) benzene 6 hours/day, 5 days/week for either 2 or 4 weeks. Of the 16 anim als 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 assessm ent of hum oral im m unocom petence using the enzym e -linked im m unosorbent assay technique. Thym us and spleen weights were determ ined, as were total cell counts for the spleen and fem u r. Spleen CD4 + /CD5 + , CD8

- /CD5 + , and Kappa + lym phocytes were determ ined by im m unostaining. Statistical evaluations were perform ed using Dunnett’s test.

Total spleen cell counts were significantly reduced (29%) in anim als exposed to 400 ppm (1278 m g /m 3 ) after 4 weeks, and thym us weight was also significantly reduced (28%). The spleen weight and cellularity were com p arable with the controls at both 2 and 4 weeks in the 200
ppm (639 m g /m 3 ) exposure group. After 4 weeks at 400 ppm (1278 m g /m 3 ) there was a significant reduction in spleen B, CD4 + /CD5

- , and CD5 + T lym phocytes. Rats exposed to 30, 200, or 400 ppm (99, 639, or 1278 m g /m 3 ) benzene for 2 or 4 weeks and then challenged with SRBCs developed a hum oral response com p arable to the controls’, and only rats exposed to 400
ppm (1278 m g /m 3 ) for 2 weeks showed a significant reduction in spleen B lym phocytes (Robinson et al., 1997). These results indicate that 400 ppm (1278 m g /m 3 ) is a LOAEL and 200
ppm (639 m g /m 3 ) is a NOAEL for im m unotoxicity in rats.

Taken as a group, the results dem onstrate that the reduction in bone m a rrow function following benzene exposure is related to decreases in the growth and m a turation of pluripotential
stem cells (CFU-S), lineage-restricted stem cells (e.g., GM-CFU), progenitor cells in various stages of m a turation, and the strom al cells that provide growth factors necessary for bone

m a rrow function. These studies tend to dem onstrate the sensitivity of erythropoiesis and lym phopoiesis to benzene and the relative resistance of granulopoiesis to benzene.

Snyder and Kalf (1994) reviewed the effects of benzene on bone m a rrow cells. Bone

m a rrow strom al cells, which include leukocytes, erythrocytes, endothelial cells, reticular cells, and fat cells, constitute the hem a topoietic connective tissue in which stem cells undergo

m a turation and am plification to form the cells of the circulation. Garnett et al. (1983) reported that the m a rrow adherent layer from benzene-treated anim als had an altered m o rphology and failed to support the differentiation of stem cells. In a related study, Kalf et al. (1996) dem onstrated that p -benzoquinone, an im portant reactive m e tabolite of benzene, prevents the conversion of pre-interleukins-1 

###  and -1 

###  to the active cytokines by inhibiting calpain and the

73interleukin-1 

###  converting enzym e , respectively. Thus, a key benzene m e tabolite inhibits the form ation and release of an im portant cytokine, IL-1, from strom al m acrophages.

Irons et al. (1992) isolated hum an CD34 + hum an bone m a rrow m y eloid progenitor cells, which proliferate in response to IL-3 and GM-CSF. Addition of hydroquinone, within a lim ited range of concentrations, resulted in synergistic enhancem ent of colony form ation. Hazel et al. (1996) m a de sim ilar observations using m ouse bone m a rrow CD32 cells. The m echanism is thought to include the up-regulation of the 5-lipoxygenase pathway in these cells, with increases in the form ation 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 m a turation of the m y eloblast stage to the

m y elocyte stage has been enhanced. These cells rem a in in their im m a ture state in the presence of hydroquinone and do not advance to m a ture neutrophils, but proliferation continues.

The im plications of these observations are significant in term s of a m e chanism by which benzene can cause expansion of the pool of m y elocytes, which do not then undergo either further differentiation to neutrophils or apoptosis, which is the usual m e thod used by the bone m a rrow to control the size of the m y elocyte pool. This m i ght provide a pathway for a benzene-induced leukem ogenic response. It m a y be hypothesized that if benzene, acting via DNA adduct form ation or oxidative dam a ge to DNA, causes cell transform ation at an earlier stage of differentiation, the creation of an expanding pool of m y elocytes by benzene and hydroquinone provides a m echanism for prom otion. Thus, the criteria for a two-stage m echanism of carcinogenesis (i.e., initiation followed by prom otion) 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 dem onstrated the production of a granulocytic leukem i a in TgAC m i ce when treated with benzene derm ally but not when given benzene by
gavage. This is the first anim al m odel that we have had for a benzene-induced granulocytic leukem i a. The TgAC m ouse is derived from the FVB m ouse in which the viral Ha-ras gene has been inserted into chrom osom e 8. The granulocytic leukem i a develops within 6 m onths. TgAC

m i ce m a y represent initiated anim als in which benzene and hydroquinone act as prom oting agents leading to proliferation of the leukem i a cells, and thus could support the argum ent for the two-stage m odel for benzene-induced leukem ogenesis. It is significant that in the course of these studies bone m a rrow depression was observed, indicating the close relationship between noncancer and cancer endpoints in benzene exposure.

The following discussion offers som e thoughts about the attem p t to understand the

m echanism by which DNA dam a ge occurs as a result of benzene exposure. DNA dam a ge is not extensively dealt with here because this docum ent concentrates on noncancer endpoints. It is not clear, however, that there is no im pact on noncancer endpoints by DNA dam a ge. Som e discussion of DNA dam a ge leading to translocation and gross breaks as well as a discussion of DNA repair and the im pact of benzene m e tabolites is warranted. The study by Snyder and Kalf (1994) and the m o re recent study by Sm ith et al. (1998) relate to benzene-induced chrom osom e dam a ge. On the issue of DNA dam a ge induced by benzene, there is am ple evidence to show that benzene m e tabolites covalently bind to protein and DNA. Also, the structures of putative DNA adducts are known. An alternative m echanism for DNA dam a ge is via the generation of reactive oxygen species. For exam ple, p -benzoquinone and hydroquinone increase superoxide, nitric

74oxide, and hydrogen peroxide in HL-60 cells activated with phorbol ester. Boersm a et al. (1994) suggested that hydroquinone itself is unlikely to generate reactive oxygen species.

An interesting com p lim entary alternative m echanism
, proposed by Singh et al. (1994), suggests that benzene- induced DNA dam a ge is m e diated by (1) release of free iron in the bone

m a rrow of benzene-treated anim als (probably by polyphenolic m e tabolites of benzene), followedby (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 dam a ge to DNA. Their suggestion that glutathionyl-hydroquinone m a y be a key interm ediate resonates well with the suggestion of Brunm ark and Cadenas (1988). In a radical-rich system , glutathionylhydroquinone m a y well be converted to glutathionyl-benzenetriol. Singh et al. (1994) suggest a chem ical 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 m a y 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 assum i ng 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 m u tagenic effect.

The discussion of the m echanism s of benzene toxicity provides a fram e work for understanding the total disease process. It has becom e clear that benzene toxicity/leukem ogenesis is a continuity of effects. Although this is not intended to be a discussion of the carcinogenic effects of benzen e, all of the effects on the bone m a rrow caused by benzene appear to be tim e- and dose-related in a sequential m a nner. The literature suggests that chronic low doses lead to cytopenias, which m a y be reversible if exposure ceases. At interm ediate chronic doses cytopenias persist and m a y eventually lead to m y elodysplastic syndrom e (MDS), which is a preleukem i c state term inating in AML. If the cytopenias are sufficiently severe the patient m a y succum b to an infection due to failure of the im m une system

. Higher doses m a y lead directly to bone m a rrow aplasia and death without the characteristic appearance of MDS. Death is usually the result of im m une deficiency. Thus, the cytopenias (the ultim ate form of which is aplastic anem ia) leading to decreases in circulating cell num bers and im m une im pairm e nt are not separable from MDS and, eventually, AML.

The following studies of exposure to benzene and its m e tabolites are considered briefly because of their value in the study of the m echanism s of benzene toxicity.

Tunek et al. (1982) exam ined the hem a totoxicity of benzene and two of its phenolic

m e tabolites, hydroquinone and catechol, in m a le NMRI m i ce. Daily doses of 440 m g /kg benzene to body weight were injected subcutaneously for 6 days, and CFU-C/tibia, cells/tibia, and m i cronuclei/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 presum e d that the num bers of m i cronuclei decreased as a result of toxicity and cell death before the cells reached the PCE stage. By day 4, the CFU-C/tibia and

num ber of cells/tibia were reduced to about 5–10% of control values.

In another experim e nt, the anim als received doses ranging from 0.7 to 440 m g /kg/day benzene on 6 consecutive days and were sacrificed on day 7. Increased levels of m i cronuclei

75were observed at doses above 28 m g /kg/day, and their frequency increased dram a tically at doses above 55 m g /kg/day. At ~100 m g /kg/day, m i cronuclei levels were about 50/2000 PCE, whereas at the highest dose of 440 m g /kg/day, they were 218/2000 PCE (vs. 3 m i cronuclei/2000 PCE for controls). CFU-C/tibia and bone m a rrow cellularity were suppressed by about 40% at doses as low as 3.5 m g /kg/day. At 100–440 m g /kg/day, both values dropped to 5–10% of control values.

Six injected doses of hydroquinone ranging from 20 to 100 m g /kg produced m i cronuclei, with a sharp increase occurring at 6 × 80 m g /kg. Bone m a rrow cellularity was slightly elevated at low doses and was suppressed at higher doses, with the num ber of CFU-C following a sim ilar pattern. Catechol, injected at doses ranging from 5 to 42 m g /kg for 6 consecutive days, had no effect on any of the three param e ters exam ined. The toxic effects of benzene were alleviated by ~30–40% when toluene (which com p etitively inhibits benzene m e tabolism ) was injected sim u ltaneously. The toxic effects of hydroquinone were som e what reduced by sim u ltaneous adm i nistration of toluene, but not nearly to the extent observed with benzene. The investigators calculated that a dose of 440 m g /kg benzene would yield an excretion of ~20 m g /kg hydroquinone. However, 20 m g /kg hydroquinone wa s virtually nontoxic, whereas 440 m g /kg benzene produced severe hem a tologic effects. On the basis of these differences and the differences in the sequence of the responses to the two chem icals, Tunek et al. (1981) speculated that the hem a totoxicity of benzene at low doses is due to agents other than hydroquinone, but that effects at and above the threshold dose result from the m e tabolic form ation and

accum u lation of hydroquinone.

Using an iron-uptake m e thod, Snyder et al. (1989) dem onstrated that i.p. adm i nistration of benzene, hydroquinone, p -benzoquinone, and MUC, singly and in com b ination with each other, inhibited erythropoiesis in fem a le m i ce. The com b ination of hydroquinone plus MUC was

m o st 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 m o rphology and function of bone m a rrow phagocytes. Male Balb/c m i ce received i.p. injections of either benzene (660 m g /kg) or a com b ination of hydroquinone and phenol (50 m g /kg each) once per day for 3 days. Control anim als received corn oil or phosphate-buffered saline. The anim als were sacrificed after the last injection and the following assays were perform ed: characterization and quantitation of subpopulations of bone m a rrow cells using m onoclonal antibody techniques,

m easurem e nt of chem otaxis (the process by which activated phagocytes m i grate to an injured site in a tissue), and m easurem e nt of the oxidative m e tabolism of phagocytes (to evaluate

m a turation and activation of phagocytes). The bone m a rrow contained three distinct populations of cells: population 1—a larger m o re dense population (41%); population 2—a population of interm ediate size and density (23%); and population 3—a sm aller, less dense population (33%). Population 1 consisted of 85–90% granulocytes, including neutrophils, basophils, and eosinophils; population 2 contained a m i xture of m ononuclear phagocytes (35–40%) and im m a ture precursor cells (55–65%); and population 3 contained lym phocytes (86%) and im m a ture precursor cells (14%). This pattern of distribution is sim ilar to that observed for hum an peripheral blood leukocytes and bone m a rrow cells (Landay and Bauer, 1988; LundJohansen et al., 1990, both cited in MacEachern et al., 1992). Differential staining of sorted

m ononuclear phagocytes revealed an increase in the num ber of m a ture, m o rphologically activated m acrophages in the bone m a rrow of benzene-treated m i ce.

76Granulocytes and m a ture m acrophages were the only cells to m i grate toward chem oattractants C5a or TPA, respectively, and cells from benzene-treated m i ce exhibited increased chem otactic activity when com p ared with controls ( p 

###  0.02). The m i gration of bone

m a rrow granulocytes from m i ce treated with hydroquinone and phenol toward C5a was depressed ( p 

###  0.02), whereas the m i gration of the m acrophages from these m i ce toward TPA was enhanced ( p <0.02).

Benzene treatm e nt also induced a significant increase in basal oxidative m e tabolism in bone m a rrow granulocytes (population 1), but treatm e nt with phenol and hydroquinone did not. The investigators concluded that phagocytes and granulocytes from the bone m a rrow of benzene-treated m i ce are activated and stim ulated to differentiate. Further studies are in progress to determ ine whether these phagocytes or their m e diators contribute to the developm ent of benzene hem a totoxicity.

Gaido and W i erda (1984) dem onstrated that in vitro treatm e nt with hydroquinone and p

benzoquinone decreased the ability of strom al cells to support granulocyte/m a crophage colony form ation. Catechol and 1,2,4-benzotriol inhibited colony form ation, but only at very high concentrations.

Eastm ond et al. (1987) dem onstrated that coadm i nistration of 75 m g /kg phenol and 25–75 m g /kg hydroquinone 2 tim es/day for 12 days i.p. to m i ce resulted in a significant, doserelated depression in bone m a rrow cellularity. The com pounds alone produced neither a significant nor a dose-related response. Catechol had no effect, either alone or in com b ination.

Hydroquinone (10 m g /kg/day), benzoquinone (2 m g /kg/day), or benzenetriol (6.25

m g /kg/day) adm i nistered i.p. to rats produced significant decreases in bone m a rrow cell counts, RBCs, and Hgb (Rao et al., 1988). Trans,trans -m uconaldehyde adm i nistered to CD-1 m i ce i.p. daily for 10 and 16 days produced hem a totoxicity sim ilar to that of benzene (W itz et al., 1985). Benzene m e tabolites added to m acrophage cultures significantly and selectively inhibited

m acrophage function (Lewis et al., 1988).

Irons et al. (1992) observed an enhanced colony-form ing response by m ouse bone

m a rrow cells treated in vitro, first with hydroquinone and then with recom b inant granulocyte/m acrophage colony-stim ulating factor. However, treatm e nt with phenol, catechol, or MUC plus the colony-stim ulating factor did not enhance colony form ation. The com b ination of hydroquinone and the stim ulating factor also appeared to recruit a segm ent of the m y eloid progenitor cell population that was not norm ally responsive to the stim ulating factor. These alterations in the differentiation of the m y eloid progenitor cell population m a y be relevant in the pathogenesis of chem ically induced AML.

4.2.1.6. Summary of Principal Hematotoxic Effects

A considerable num ber of studies in experim e ntal anim als have pointed to the changes in peripheral blood and bone m a rrow induced by benzene as being am ong the m o st sensitive effects of the com pound’s toxicity. That these effects were seen in m o st, if not all, of the species tested, as well as in hum ans occupationally exposed to benzene (see Section 4.1.2.1.), suggests that their threshold dosim etry m a y be im portant in identifying a point of departure for the derivation of

77quantitative risk estim ates. Principal am ong the adverse effects described are leukopenia, lym phocytopenia, granulocytosis, anem ia, and reticulocytosis. Com paring the incidence of one response with another, lym phocyte counts appear to be depressed sooner and m o re severely than other cell types, and granulocytes m a y be the m o st resistant (Snyder et al., 1978, 1980). Differences in responsiveness to benzene have been observed am ong species, with m i ce being

m o re sensitive than rats to the hem a totoxicity of benzene (W ard et al., 1985; Snyder et al., 1978, 1980). Intraspecies variability was dem onstrated by Snyder et al. (1978, 1980), who reported that AKR/J m i ce were m o re sensitive than C57BL/6 m i ce to the com pound. W a rd et al. (1985) showed that in CD-1 m i ce, m a les were m o re sensitive than fem a les.

A striking feature of the studies of W a rd et al. (1985) and Cronkite et al. (1985) was that hem a tologic abnorm alities were evident after only 2 weeks of exposure, with com p aratively 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 elem ents m a y have the capacity to recover from benzene-induced toxicity, depending on the duration of exposure and concentration. For

exam ple, lym phocyte counts, which were depressed in C57BL/6 m i ce exposed to benzene at 300
ppm (32–1279 m g /m 3 ) for 2–16 weeks, tended toward norm al values during a 4- to 8-week recovery period.

Dose-response studies such as that of W a rd et al. (1985) have served as a basis for the identification of NOAELs and LOAELs. For exam ple, a NOAEL of 30 ppm (96 m g /m 3 ) and a LOAEL of 300 ppm (958 m g/m 3 ) have em erged for CD-1 m i ce from the W a rd et al. study. The data from this study have also served as a basis for BMD m odeling. Using the BMDL derived from this m odeling as a point of departure for RfC determ ination yields a value that is in good
agreem ent with the RfC obtained from the hem a totoxicity data from the occupational exposure study of Rothm an et al. (1996a), as described in Section 5.1.

Benzene-induced peripheral blood abnorm alities reflect a disruption of all levels of hem a topoiesis in the bone m a rrow. For exam ple, several studies have dem onstrated concentration-dependent depression of bone m a rrow cellularity in m i ce treated with benzene, the lowest concentration producing this effect being 21 ppm adm i nistered for 2 weeks (Toft et al., 1982). Sim ilarly, the stem cell com p artm ent (CFU-S) appears to be sensitive to the adverse effects of benzene, particularly when drawn into the cell cycle in response to benzene-induced dam a ge to other cells (Snyder, 1987). Thus, concentrations of 100 ppm (320 m g /m 3 ) benzene have produced significant depressions in the CFU-S population in the bone m a rrow of m i ce 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 com pound, exhibiting significant depression at 21 ppm (67 m g /m 3 ) benzene for 2 weeks and 10 ppm (32 m g /m 3 ) 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 m a rrow m acrophages also appear to be a target of benzene toxicity (MacEachern et al., 1992). Thus, strom al m acrophages from benzene- or hydroquinone-treated m i ce failed to

78convert IL-1 to the m a ture cytokine 17-Kd, a com ponent essential for hem a topoiesis (Renz and Kalf, 1991). This allows the suggestion that the inhibitory effect of benzene on lym phocyte proliferation can be m e diated through the suppression of cytokine production.

Another im portant consideration is whether the benzene-induced perturbation of the hem a topoietic system is m e diated through the effects of the parent com pound or via one or m o re

m e tabolites. The fact that both in vivo and in vitro studies have dem onstrated the ability of som e

m e tabolites to induce qualitatively sim ilar toxicological effects to those of benzene provides evidence consistent with the concept that the hem a totoxicity of the com pound m a y be m e diated through its m e tabolites.

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 inform ation pertinent to the reproductive toxicity of benzene in anim als 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 m i ce and rats (NTP, 1986). Other than ovarian effects, no signs of reproductive toxicity were reported. Benzene was adm i nistered by gavage to B6C3F1 m i ce and F344/N rats. In the 90-day study, the doses of benzene, adm i nistered in corn oil 5 days/week, ranged from 25 to 600 m g /kg; in the 2-year study, doses were 50, 100, or 200 m g /kg for m a le rats and 25, 50, or 100 m g /kg for fem a le rats and all m i ce. There were no ovarian effects in either the rats or the m i ce 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 num ber of anim als exam ined for each group, for m i ce adm i nistered 0, 50, 100, or 200 m g /kg benzene were ovarian atrophy (32, 79, 65, and 46%, respectively) and epithelial hyperplasia (26, 89, 63, and 60%, respectively).

Spano et al. (1989) exam ined the cytotoxic action of benzene on m ouse germ cells using flow cytom e tric DNA content m easurem e nts. Testicular m onocellular suspensions were obtained from (C57BL/Cne × C3H/Cne) F1 m a le m i ce receiving single doses by gavage of 0, 1, 2, 4, 6, or 7 m L /kg benzene to body weight. The effects were m easured in three anim als per group 7, 14, 21, 28, and 70 days after treatm e nt. Testicular cells were classified as m a ture haploid, im m a ture haploid (haploid is split into two parts because of different staining on the elongated and round sperm a tids), diploid, or tetraploid, depending on DNA content.

Benzene treatm e nt did not affect body weight or testes weight, but it did alter the ratio of testicular cell types. DNA histogram s of m ouse testis cells obtained at different tim es after benzene exposures showed a dose-related decrease in the tetraploid cell fraction (m ainly prim a ry sperm a tocytes); 7 days after treatm e nt the tetraploid cell num ber in anim als exposed to 6 and 7

m L /kg benzene was depressed to ~80% of control values. The percentage of round sperm a tids in anim als exposed to 7 m L /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

79differentiating sperm a togonia. Dose-dependent recovery processes for the tetraploid cells had begun 21 days after treatm e nt, sim u ltaneously with reduction of other cell subpopulations, but they were still incom p lete by 70 days. Recovery of the round sperm a tids began by day 28 posttreatm e nt and was com p lete by day 70. There was also a tim e- and exposure-dependent reduction (to 60% of control values for the 4, 6, and 7 m L /kg doses) in the percentage of elongated sperm a tids at 28 days after treatm e nt; recovery was com p lete by 70 days. The investigators concluded from this study that benzene can induce acute cytotoxicity in m ouse germ cells. Both the NTP (1986) and the Spano et al. (1989) studies provide valuable descriptive inform ation regarding the effects of benzene on the reproductive system

. It should be noted that longer term and m o re 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 anim al species is sum m a rized in Table 7.

Subchronic toxicity studies have provided inform ation relevant to the reproductive toxicity of benzene. In an early inhalation study, W o lf et al. (1956) dem onstrated m oderately increased testicular weights in groups of 10–25 rats exposed to 6600 ppm (21,084 m g /m 3 ) benzene for 13 weeks but not in rats exposed to 88 ppm (281 m g /m 3 ) for 30 weeks, to 2200 ppm (7028 m g /m 3 ) for 30 weeks, to 4400 ppm (14,056 m g /m 3 ) for 5 weeks, or to 9400 ppm (30,030
mg / m 3 ) for m o re than 1–19 days. Groups of 5–10 m a le guinea pigs exposed to 88 ppm (281
mg / m 3 ) benzene for ~9.6 m onths had a slight increase in testicular weights, but guinea pigs

Table 7. Reproductive toxicity of inhaled benz ene in test animals

Species No./
group Treatment Reproductive effects

Male rat 10–25 21,084 m g /m 3 , 7 hrs/day, 5 days/wk for 13 weeks Moderate changes in testes weight (no quantitative data)

Rabbit

Male guinea pig

Male and fem a leSpragueDawley
rats LOAEL/ NOAEL (mg/m 3 ) R eference

LOAEL: NA NOAEL: NA W o lf et al., 1956

LOAEL: NA NOAEL: NA W o lf et al., 1956

LOAEL: NA NOAEL: NA W o lf et al., 1956 256 m g/m 3 , 7 hrs/day, 5 day/wk for ~8.5 m onths

5–10 281 m g/m 3 , 7 hrs/day, 5 days/wk for ~9.6 m onths

50/sex 0 , 3.2, 32, 96, 958
mg / m 3 6 hrs/day, 5 days/wk for 13 weeks; 10/sex/group sacrificed after 7, 14, 28, 56, and 91 days of treatm e nt 1–2 Slight change in histopathology of the testes (no details)
Slight change in weight of testes (no quantitative data)

No m i croscopic lesions of ovaries and testes LOAEL: 958 a NOAEL: 96 W a rd et al., 1985

80Species No./
group Treatment Reproductive effects LOAEL/ NOAEL (mg/m 3 ) R eference

CD-1

m ouse 150/sex 0 , 3.2, 32, 96, 958
mg / m 3 6 hrs/day, 5 days/wk for 13 weeks; 30/sex/group sacrificed after 7, 14, 28, 56, and 91 days of treatm e nt Cystic ovaries (4/10), testicular atrophy (7/10), decreased sperm count (6/10), and increased abnorm a l sperm form s (9/10) only at 958 m g/m 3 for 91 days LOAEL: 958 NOAEL: 96 W a rd et al., 1985

Fem a leSpragueDawley rat 26 0, 3.2, 32, 96, 958
mg / m 3 6 hrs/day, 5 days/wk for a 10- week period before and during m a ting, during gestation to GD 20, and from day 5 to day 21 of lactation Reduction in body and liver weight in the fem a le pups at 958 m g/m 3 ( p <0.05); no effect on m a ternal

m ortality, body weight, physical param e ters, pregnancy rate, length of gestation, num ber of live and dead pups at birth, and sex distribution data; no effect on pup survival and growth, and gross postm ortem m a nifestations LOAEL: 958 NOAEL: 96 Kuna et al., 1992

a Effects on hem a tologic param e ters place the LOAEL for the overall study in rats at 958 m g /m 3 and the NOAEL at 96 m g/m 3

. GD = gestation day. NA = not applicable (only one concentration tested).

81exposed to 88 ppm (281 m g /m 3 ) for 4 weeks did not. Groups of 1–2 rabbits exposed to benzene concentrations of 88 ppm (281 m g /m 3 ) for ~8.5 m onths had slight histopathologic changes in the testes, described by the authors as degeneration of the sem i niferous tubules (no other concentrations were tested in rabbits). Exposures were routinely 7 to 8 hours/day, 5 days/week. Air-exposed or unexposed anim als served as controls. The vapor concentrations were

m a intained 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 stim ulated further interest in the potential reproductive effects of benzene.

W a rd et al. (1985) exposed m a le and fem a le CD-1 m i ce and Sprague-Dawley rats to benzene concentrations of 1, 10, 30, or 300 ppm (3.2, 32, 96, or 958 m g /m 3 ) 6 hours/day, 5 days/week for 13 weeks. Anim als exposed to filtered air served as controls. Groups of 10 m a les and 10 fem a les per species were sacrificed after 7, 14, 28, 56, and 91 days of treatm e nt. Endpoints of the investigation that were relevant to reproductive toxicity included body and organ weights and gross and m i croscopic pathology. The m i ce exposed to 300 ppm (958 m g /m 3 ) 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 com p arable with control values. The m i ce exposed to 300 ppm (958 m g /m 3 ) for 91 days had testicular lesions that included m i nim a l to m oderately severe bilateral atrophy and degeneration (7/10 m i ce), m oderate to m oderately severe decreases in sperm a tozoa in the epididym al ducts (6/10 m i ce), and m i nim a l to m oderate increases in abnorm al sperm form s (9/10). Four of 10 fem a le m i ce had bilateral ovarian cysts after 91 days of exposure to 300 ppm (958 m g /m 3 ) benzene. These testicular and ovarian lesions were not observed at earlier sacrifices, but sim ilar lesions did appear at lower concentrations; however, biological significance is questionable. Microscopic findings in the control anim als were not m e ntioned.

The rats had no signs of reproductive toxicity; however, hem a totoxicity occurred at 300
ppm (958 m g /m 3 ), giving the overall study a LOAEL of 300 ppm (958 m g /m 3 ) and a NOAEL of 30 ppm (96 m g /m 3 ). The experim e nt appeared to be a carefully perform ed subchronic toxicity study, and the appearance of m i croscopic changes in the testes and ovaries of the 300 ppm (958
mg / m 3 ) group and sim ilar—but only occasional—fi ndings at 10 and 30 ppm groups suggest a concentration response.

Kuna et al. (1992) assessed the effects of benzene on fem a le fertility in Sprague-Dawley
rats. The study tested occupational exposure levels as well as higher exposures that previously dem onstrated developm ental toxicity. Groups of 26 Sprague-Dawley fem a le rats inhaled benzene vapor concentrations of 0, 1, 30, and 300 ppm (0, 3.2, 96, and 958 m g /m 3 ) benzene (purity, 99.96%) 6 hours/day, 5 days/week for a 10-week prem a ting and m a ting period, and daily from gestation days (GDs) 0 to 20 and lactation days 5 to 20. Daily vaginal sm ears were

exam ined to determ ine whether estrus was affected by treatm e nt.

Neither strain had any treatm e nt-related effects, as evidenced by data on m a ternal

m o rtality, body weight, physical param e ters, pregnancy rate, length of gestation, num ber of live and dead pups at birth, and sex distribution. No effect was observed on pup survival and growth or on gross postm ortem 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 m g /m 3 )

82levels; 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 fem a le pups at 300
ppm (958 m g /m 3 ). Increases in relative kidney weight for the fem a le 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 m g /m 3 ) 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 m g /m 3 ) and a NOAEL of 30 ppm (96 m g/m 3 ), based on the statistically significant reduction in body and liver weights, and based on past studies showing sim ilar effects at sim ilar 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 chem icals of known and unknown developm ental toxicity potential, to validate the Chernoff/Kavlock assay as a screen for developm ental effects. Benzene doses of 1300

m g /kg/day were adm i nistered in cottonseed oil by gavage to pregnant ICR/SIM m i ce on GDs 8 through 12, and the dam s were allowed to deliver. Maternal toxicity was assessed using body weights, m o rtality, or other clinical signs of overt toxicity as endpoints. Benzene had no effect

on m a ternal body weight, but it did produce significantly lower neonatal body weight (to ~95% of control value, p <0.05), as m easured 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 inform ation. However, the study does dem onstrate evidence for fetal toxicity in the absence of observable m a ternal toxicity.

In the Exxon Chem i cal Com pany (1986) study, pregnant fem a le Sprague-Dawley rats (20–22/group) were treated by gavage with 0, 50, 250, 500, or 1000 m g /kg/day benzene on GDs 6–15. No dose-related m o rtality was observed. Significant findings in the treated dam s as

com p ared with controls were decreased food consum ption at 

###  250 m g /kg, decreased body weight, body weight gains at 

###  500 m g /kg, and increased incidence of alopecia at 1000 m g /kg ( p <0.05). Developm ental toxicity was lim ited to decreased fetal body weights at 

###  500 m g /kg ( p <0.05). Fetuses were exam ined only for external m a lform ations, not for skeletal and visceral

m a lform ations. This study identified a NOAEL of 50 m g /kg/day and a LOAEL of 250

m g /kg/day for m a ternal toxicity and a NOAEL of 250 m g /kg/day and a LOAEL of 500

m g /kg/day for developm ental toxicity in Sprague-Dawley rats.

4.2.2.2.2. Inhalation exposure

. This section sum m a rizes the data for the developm ental toxicity of inhaled benzene in anim als. The studies included experim e nts with three species (rats, m i ce, and rabbits) and interm ittent (6 or 7 hours/day) and continuous (24 hours/day) exposures. Som e studies tested m o re than one concentration of benzene and dem onstrated a concentration response and a LOAEL and/or a NOAEL. Som e studies used only one concentration and were not useful for determ ining dose-response relationships, but they did provide supporting evidence for fetal toxicity. Three studies evaluated fetal hem a tologic param e ters, and one study exam ined the possible synergistic effects of benzene with other chem icals. The available studies on the developm ental toxicity of inhaled benzene are sum m a rized in Table 8.

83Table 8.Developmental toxicity of inhaled benzene in test animalsStrain/speciesNo. ofdams/groupExposureaEffectsEffect levelsmaternal/developmental(mg/m3)ReferenceMaternalDevelopmentalSprague-14–150, 32, 160, or 1597Decreased bodyDecreased mean live bodyLOAEL: 160/160Kuna andDawley ratexposed,mg/m3, 7 hrs/day onweight and bodyweight at 160 and 1597NOAEL:32/32Kapp, 198111 controlsGDs 6–15; sacrificedweight gain at 160mg/m3, day 20c; decreasedon day 20and 1597 mg/m3,crown-rump distance, 1597days 5–15; dose-mg/m3c; skeletal andrelatedbvisceral (brain)abnormalities at 160 and1597 mg/m3; increasedincidence of malformationsat 1597 mg/m3SpragueDawley rat35–37exposed,32–34controls0, 3.2, 32, 128, or319 mg/m3, 6 hrs/dayon GDs 6–15;sacrificed on day 20None in any groupDecreased body weight at319 mg/m3(p0.05);variants in all but one group(including controls), notdose related; no increase inincidence of malformationsLOAEL:NA/319NOAEL:NA/128Coate etal., 1984SpragueDawley rat14–180, 319, 958, or 7028mg/m3, 6 hrs/day onGDs 6–15; sacrificedon day 21Decreased bodyweight gain at 7028mg/m3(p< 0.01)Decreased body weight andlength at 7028 mg/m3(p<0.05); increased skeletalvariants all exposure groups(p<0.05 at 319 and 7028mg/m3; females moresensitive than males); noincrease in incidence ofmalformationsLOAEL: 7028/319NOAEL: 958/NAGreen etal., 1978Rat5–121.0–670 mg/m324hrs/day, 10–15 daysbefore mating andthroughoutpregnancyNo dataTendency toward decreasedlitter sizes at 64 mg/m3;complete absence of littersat 670 mg/m3LOAEL:NDNOAEL:NDGofmekler,1968

84Table 8.Developmental toxicity of inhaled benzene in test animals(continued)Strain/speciesNo. ofdams/groupExposureaEffectsEffect levelsmaternal/developmental(mg/m3)ReferenceMaternalDevelopmentalCFY rat19 exposed,28 controls0 or 1000 mg/m3, 24hrs/day on days9–14 of pregnancyDecreased bodyweight gain (p<0.01)Decreased body weight(p<0.01) and growthretardation (p<0.05);retarded skeletaldevelopment and increasedincidence of extra ribs andfused sternebrae (p<0.05for both); no increase inincidence of malformationsLOAEL:NANOAEL:NAHudak andUngvary,1978CFY rat20–22exposed,48 controls0, 150, 450, 1500, or3000 mg/m3, 24hrs/day on GDs7–14; sacrificed onday 21Decreased bodyweight gain at150mg/m3(p<0.001),somewhat doserelated; liver/bodyweight increased(p<0.05 or 0.01)Decreased body weight at150 mg/m3(p<0.001),increased resorptions andskeletal and weightretardation (p<0.01–0.05),effects not dose related; noincrease in incidence ofmalformationsLOAEL: 150/150NOAEL:NATatrai etal., 1980CFY rat17400 mg/m3, 24hrs/day on GDs 7–15Decreased bodyweight gain(p<0.001); increasedrelative liver weight(p<0.05)Retarded weight gain(p<0.01); skeletal growthretardation (p<0.001)LOAEL:NANOAEL:NAUngvary,1985

85Table 8.Developmental toxicity of inhaled benzene in test animals(continued)ReferenceKeller andSnyder,1988Corti andSnyder,1996Effect levelsmaternal/developmental(mg/m3)LOAEL:NA/64NOAEL: 64/32LOAEL:NANOAEL:NAEffectsDevelopmental16-day-old fetus: no effecton hematologic parameters2-day-old neonates:reduced circulatingerythroid precursor cells(all concentrations) (p<0.05at 64 mg/m3); increasedhepatic hemato-poietic blastcells, lympho-cytes, andgranulopoietic precursorcells and decreased hepaticerythropoiesis precursorcells (allp<0.05 at 64mg/m3)6-week-old adult:similarpattern of enhancedgranulopoiesis (64 mg/m3)Bone marrow samples from6-week-old offspring:protocols (a) and (b) causedchanges in CFU-E counts,males only; protocol (c)caused changes in CFU-Ecounts, females onlydMaternalNone observedNo dataExposurea16, 32, or 64 mg/m3,6 hrs/day on GDs6–15(a) 32 mg/m3benzene GDs 6–15(b) 5% ethanol indrinking water ad lib(c) 32 mg/m3benzene + 5%ethanol(d) air + distilledwaterNo. ofdams/group58Strain/speciesSwiss-WebstermouseSwiss-Webstermouse

86Table 8.Developmental toxicity of inhaled benzene in test animals(continued)Strain/speciesNo. ofdams/groupExposureaEffectsEffect levelsmaternal/developmental(mg/m3)ReferenceMaternalDevelopmentalCF-1 mouse35–370 or 1597 mg/m3ppm, 7 hr/day onGDs 6–15; sacrificedon day 18NoneDecreased body weight(p<0.05), “significantly”increased skeletal variantsof fetuses; no increase inincidence of malformationsbut was toxic to fetusesLOAEL:NANOAEL:NAMurray etal., 1979CFLP mouse15 exposed,115 controls0, 500 or 1000mg/m3, 24 hrs/day onGDs 6–15Not mentionedWeight and skeletalretardation, bothconcentrations (p< 0.05),somewhat dose-relatedLOAEL:ND/500NOAEL:NA/NAUngvaryand Tatrai,1985New Zealandrabbit200 or 1597 mg/m3, 7hr/day on GDs 6–18;sacrificed on day 29NoneStatistically significantdecrease in minor skeletalvariants, lumbar spurs andproportion with 13 ribs, inexposed fetusesLOAEL:NANOAEL:NAMurray etal., 1979New Zealandrabbit11 or 15exposed,60 controls0, 500 or 1000mg/m3, 24 hrs/day onGDs 7–20Decreased weightgain and increasedrelative liver weightat 1000 mg/m3(p<0.05)Decreased body weightand increased abortions andskeletal variants at 313 ppm(p<0.05 for all effects)LOAEL: 1,000/1,000NOAEL: 500/500Ungvaryand Tatrai,1985aConversion factors, 1 ppm = 3.26 mg/m3; 1 mg/m3= 0.31 ppmbStatisticallydifferent fromcontrol as determined bypairwise multiple comparison procedurescStatisticallydifferent from control as determined byCochran’s approximation tot (t’)dData fromabstract; no other details availableGD = gestation dayNA = not applicableND = not determined

87In a review of som e of the earlier, m o stly unpublished developm ental toxicity studies on benzene, Brief et al. (1980) observed that the developm ental effects in anim als were characterized m a inly by fetal toxicity in rats exposed to 40, 50, 500, and 2200 ppm (128, 159, 1597, and 7028 m g /m 3 ) (Dow Chem i cal Co., 1992b; Green et al., 1978). Maternal toxicity and

som e fetal m a lform ations occurred at 500 ppm (1597 m g /m 3 ); concentrations of 10 ppm (32
mg / m 3 ) 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 m g /m 3 ) 7 hours/day on GDs 6–15 and sacrificed the dam s on day 20 to evaluate m a ternal and developm ental effects. Hem a tologic evaluation of the dam s , perform ed on GDs 5 and 20, included RBC, W BC, and differential
counts.

There were no deaths, observable illness, or hem a tologic changes for dam s in any
exposure group. During GDs 5 through 15, m a ternal body weight gains were significantly decreased to 66% of control values in the 50 ppm (160 m g /m 3 ) group and to 63% of control values in the 500 ppm (1597 m g /m 3 ) group during GDs 5–15 ( p <0.05 for both groups), whereas during GDs 15–20, body weights in the dam s exposed to 10 ppm (32 m g /m 3 ) and weight gain in the dam s exposed to 10 and 500 ppm (32 and 1597 m g /m 3 ) were increased. Body weight corrected for gravid uterine weight was not determ ined. No differences were observed between exposed and control groups in the num ber of im plantation sites per num ber of ovarian corpora lutea (im p lantation efficiency); the incidences of resorbed, dead, or live fetuses; and the sex distribution.

Fetal-crown rum p lengths were decreased in the 500 ppm (1597 m g /m 3 ) group, and m ean body weight of live fetuses were decreased in both the 50 and 500 ppm (160 and 1597 m g /m 3 ) groups (control, 4.4 ± 0.6 g; 50 ppm [160 m g /m 3 ], 3.8 ± 0.7 g; 500 ppm [1597 m g /m 3 ], 3.6 ± 0.8 g; all statistically significant, p <0.05). One fetus from each of four litters from dam s treated with 500 ppm (1597 m g /m 3 ) benzene displayed exencephaly (one fetus), angulated ribs (one fetus), or ossification of the forefeet out of sequence (two fetuses); these abnorm alities were not observed in control fetuses. This group also exhibited delayed ossification in the skull, vertebral colum n , rib cage, pelvic girdle, and extrem ities 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

m ean num ber of phalanges and m e tacarpals.

In sum m a ry, 13 litters and 142 fetuses were exam ined from the group exposed to 500
ppm (1597 m g /m 3 ); 30 fetuses from six litters had delayed ossification ( p <0.05), and four fetuses from four litters had skeletal variants and abnorm alities. In the group exposed to 50 ppm (160
mg / m 3 ), 125 fetuses from 15 litters were exam ined; 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 m g /m 3 ). The study appears to have been conducted according to standard protocols and recom m e ndations (U.S. EPA, 1991) for developm ental toxicity. The study establishes a LOAEL for the developm ental toxicity of inhaled benzene of 50 ppm (160 m g /m 3 ) and a NOAEL of 10 ppm (32 m g /m 3 ).

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 m g /m 3 ) benzene 7 hours/day on GDs 6–15. Benzene levels

88in the cham bers, m onitored at least once per day using both infrared and gas chrom atography

m e thods, were within 10% of target concentrations by both analyses, except for the gas chrom atography analysis of the 10 ppm (32 m g /m 3 ) concentration, which was within 17% of target.

Maternal body weight and m o rtality rate did not differ between treated and control anim als (other than a slight but insignificant decrease in body weight of the 100 ppm [319
mg / m 3 ] group), and there were no clinical signs or gross pathology suggestive of m a ternal toxicity during gestation. Body weight corrected for gravid uterine weight was not determ ined, and hem a tologic assays were not perform ed. The average num ber of im plantations, num ber of resorptions, resorption incidence, num ber of live fetuses, and the incidence of dam s with one or

m o re resorbed im plantations were com p arable in all control and exposed groups. The fetuses in the 100 ppm (319 m g /m 3 ) group exhibited a significant decrease ( p 

###  0.05) in average m a le and fem a le 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 m o re variants, although slight increases in variants (dilation of the renal pelvis and ureters) occurred in the 1 and 100 ppm (3.2 and 319 m g /m 3 ) groups. Soft tissue

anom alies occurred in all exposed groups, but they were not dose related. Delays in ossification of the skull, vertebral centra, and extrem ities occurred at 100 ppm (319 m g /m 3 ). The investigators concluded that benzene is weakly toxic to the fetus at 100 ppm (319 m g /m 3 ) (a concentration that was not toxic to the dam s ) and is not toxic to the fetus at 40 ppm (128 m g /m 3 ). This study establishes a LOAEL of 100 ppm (319 m g /m 3 ) and a NOAEL of 40 ppm (128 m g /m 3 ) 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 m g /m 3 ) 6 hours/day on GDs 6–15. Daily body weights were m easured as an indicator of m a ternal toxicity; peripheral blood cell counts were not perform ed. The dam s were sacrificed on day 21, and developm ental toxicity was evaluated. Maternal weight gain was sim ilar to control values for the rats exposed to 100 and 300 ppm (319 and 958 m g /m 3 ); however, body weight for the fem a les exposed to 2200 ppm (7028 m g /m 3 ) 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 im plantation sites/litter, live fetuses/litter, percentage resorption/im plantation site, percentage litters with resorption, litters totally resorbed, and resorption/litter with resorption. At concentrations of 100 and 300 ppm (319 and 958 m g /m 3 ), fetal sex ratio, m ean fetal body weight, and m ean fetal crown-rum p length were com p arable with control values; at 2200 ppm (7028 m g /m 3 ), m ean fetal weight and m ean crown-rum p length were significantly lower than controls ( p <0.05).

The incidence of delayed ossification was sim ilar for the exposed groups and controls; however, am ong the fetuses exposed to 300 and 2200 ppm (958 and 7028 m g /m 3 ) and exhibiting delayed ossification, the incidence was increased significantly for fem a les over m a les ( p <0.05). The num ber of litters displaying an increased incidence of unossified sternebrae was increased at 100 ppm (319 m g /m 3 ) (9/18 litters) and 2200 ppm (7028 m g /m 3 ) (11/15 litters), and at 2200 ppm (7028 m g /m 3 ) the num ber of fem a les with the abnorm ality was significantly increased over

ma l e s ( p <0.05). The concentration of 2200 ppm (7028 m g /m 3 ) benzene was m a ternally toxic; however, fetal toxicity m a nifested as skeletal abnorm alities was observed at concentrations that

89were not m a ternally toxic (and were statistically significant). The apparently increased sensitivity of the fem a le fetus to the effects of benzene is supported by other observations that fem a le rabbits, m i ce, and rats are m o re 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 m a les and fem a les m a y be related to horm onal differences.

In an early study, Gofm ekler (1968) exposed fem a le rats to benzene concentrations ranging from 0 to 210 ppm (670 m g /m 3 ) 24 hours/day for 10–15 days prior to m a ting 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 m g /m 3 ). This effect was not observed in the experim e nts of Green et al. (1978) at higher benzene concentrations adm i nistered over a different exposure period (6 hours/day on GDs 6–15). This suggests that the rat em bryo is m o re sensitive to continuous exposure than to interm ittent exposure only during organogenesis. However, fem a les in that study were exposed before m a ting, and ovulation rate and fertility m a y have been affected independently of developm ental effects. The Gofm ekler (1968) study dem onstrated significant differences in the weight of liver, lung, spleen, kidney, and thym us from fetuses of benzene-treated dam s

. The weights were m o stly decreased, but the effect was not dose related. Gofm ekler (1968) concluded that exposure of pregnant rats to relatively low concentrations of benzene can disturb norm al fetal developm ent. However, the lack of a concentration response in the organ weight data for fetal rats precludes the determ ination of a LOAEL or a NOAEL for the study.

Hudak and Ungvary (1978) conducted an experim e nt with CFY rats exposed to 313 ppm (1000 m g /m 3 ) benzene 24 hours/day on GDs 9 through 14. Exposure to benzene resulted in statistically significant decreases in m a ternal 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) exam ined the developm ental toxicity of benzene using several doses. Pregnant CFY rats inhaled 0, 47, 140, 465, or 930 ppm (0, 150, 450, 1500, or 3000 m g /m 3 ) 24 hours/day on GDs 7–14. Anim als were sacrificed on day 21 to evaluate developm ental effects. The dam s 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 m g /m 3 ), but not at 930 ppm (3000 m g /m 3 ). For exam ple, the body weight gain for dam s exposed to 465 ppm (1500 m g /m 3 ) benzene was 28.5% of the starting body weight

com p ared with the control value of 62.8%, whereas the body weight gain for dam s exposed to 930 ppm (3000 m g /m 3 ) benzene was 37.0%. Liver/body weight ratios were significantly increased in the dam s exposed to 140, 465, or 930 ppm (450, 1500, or 3000 m g /m 3 ), with a response pattern sim ilar to the body weights ( p <0.05 or p <0.01, ANOVA).

Fetal body weights were decreased at 

###  47 ppm (150 m g /m 3 ) ( p <0.001, all concentrations); fetal loss in the percent of total im plantation sites was increased 4–7 tim es control values at 

###  140 ppm (450 m g /m 3 ) ( p <0.05 or p <0.01, Mann W h itney 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 m a lform ations, even at

90concentrations 

###  47 ppm (150 m g /m 3 ) that were m a ternally toxic. The m a in difference observed for the days 7–14 exposure regim e n (Tatrai et al., 1980) study versus the days 9–14 regim e n (Hudak and Ungvary, 1978) was the increase in fetal m o rtality (30–40% of the im plants) that occurred when exposure was started on GD 7.

To test the hypothesis that the congenital effects of industrial solvents m a y be attributable to sim u ltaneous exposure to a com b ination of two or m o re chem icals, Ungvary and Tatrai (1985) conducted a study using com b inations of benzene and its m e thyl derivatives, toluene and xylene, with each other or with acetylsalicylic acid (ASA), a known developm ental toxicant. Groups of pregnant CFY rats inhaled com b inations of 123 ppm (400 m g /m 3 ) benzene and 307 ppm (1000
mg / m 3 ) toluene or 189 ppm (600 m g /m 3 ) xylene 24 hours/day on GDs 7–15. Additional groups of 7–14 anim als were exposed to 798 ppm (2600 m g /m 3 ) benzene on GDs 10–12 and given oral doses of either 250 or 500 m g/kg of ASA on GD 12.

At 400 m g /m 3 , benzene produced m a ternal toxicity, as evidenced by decreased weight gain (47% gain in the benzene-treated dam s 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 anom alies or skeletal, internal, or external m a lform ations or the percentage of abnorm al survivors. The com b ined developm ental effects of the solvents were not additive. However, when each solvent was adm i nistered by inhalation in com b ination with orally adm i nistered ASA, it enhanced m a ternal as well as fetal toxicity. In addition, benzene and its m e thyl derivatives significantly increased the frequency of ASA-induced m a lform ations. The induction of m o st

m a lform ations was not solvent dependent; cleft lip and palate and abnorm alities of the spinal colum n occurred only in response to the com b ined treatm e nt with solvents and ASA. The investigators concluded that under the conditions of this experim e nt, neither benzene nor its alkyl derivatives induced m a lform ations, alone or in com b ination, and that the fetal toxicity observed with all three solvents was not additive when the solvents were com b ined; however, all three solvents potentiated the toxic effects of ASA, including the induction of m a lform ations.

Pregnant Swiss-W e bster m i ce (five per exposure level per progeny age group; initial age 8–12 weeks) were exposed via inhalation to nom inal vapor concentrations of 0, 5, 10, or 20 ppm (0, 16, 32, or 64 m g /m 3 ) 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 m easure hem a topoietic progenitor cells (CFU-E, BFU-E, and GM-CFU-C) from the liver (fetuses and neonates) and bone m a rrow and spleen (adults). In addition, 10-week-old progeny from litters in the control and m i d-exposure group were exposed for 2 weeks to 10 ppm (32 m g /m 3 ) benzene, then sacrificed to m easure hem a topoietic progenitor cells from the bone m a rrow and spleen.

There was no evidence of m a ternal or nonhem a topoietic developm ental toxicity in benzene-exposed m i ce. There was a significant increase ( p <0.05) in the num bers of BFU-E from livers of m a le and fem a le fetuses exposed to the low- and m i d-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 m i d-exposure levels and decreases at the highexposure level; in m a le neonates, there were increases and decreases in liver CFU-E at the m i dexposure level and increases at the high-exposure level; and in adult m i ce, there were decreases in bone m a rrow CFU-E and increases in spleen CFU-E in m a les exposed to 10 ppm (32 m g /m 3 )

91in utero. Liver GM-CFU-C in neonates was significantly decreased ( p <0.05) at the m i dexposure level (m ales only) and increased at the high-exposure level. Mice exposed to 10 ppm (32 m g /m 3 ) benzene in utero and for 2 weeks as adults had significantly ( p <0.05) decreased bone

m a rrow CFU-E (m ales only) and splenic GM-CFU-C; m i ce exposed to air in utero and 10 ppm (32 m g /m 3 ) benzene for 2 weeks as adults had no changes in bone m a rrow or splenic CFU-E, but they had a significant ( p <0.05) decrease in splenic GM-CFU-C (fem a les only). The authors concluded that benzene treatm e nt in utero induced hem a topoietic 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 m g /m 3 ) for developm ental hem a topoietic effects in m i ce, because statistically significant changes in BFU-E and CFU-E were observed in the 16-day fetuses from dam s exposed to 5 ppm (16 m g /m 3 ). 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 m g /m 3 ) followed by decreases at 20 ppm (64 m g /m 3 ). Only five pregnant anim als were used in the study per exposure dose level, and lim ited num bers of anim als (two fetuses, two neonates, or one adult/litter/sex) were used for the evaluation of hem a totoxic effects. Also, the responses did not establish a consistent pattern in the different ages of progeny exam ined. Therefore, there is a high degree of uncertainty associated with determ ining whether the effects observed at 5 ppm (16 m g /m 3 ) are truly adverse effects.

Keller and Snyder (1988) further exam ined the effects of in utero b enzene on the developing recognizable hem a topoietic precursor cells and peripheral blood cells and Hgb production in Swiss-W e bster m i ce. Three separate exposure experim e nts were perform ed in which pregnant m i ce (five per group) were exposed to benzene concentrations of 0, 5, 10, or 20
ppm (0, 16, 32, or 64 m g /m 3 ) in cham bers 6 hours/day on GDs 6–15. Maternal toxicity was evaluated on the basis of m o rbidity, m o rtality, or weight loss, but m a ternal peripheral blood count was not determ ined. In experim e nt 1, the fetuses of the benzene-exposed dam s were assayed for signs of hem a totoxicity on GD 16 (two fetuses/sex/litter). In experim e nt 2, the offspring (two neonates/sex/litter) were assayed at 2 days. In experim e nt 3, progeny (one/sex/litter) were exam ined at 6 weeks of age. Param e ters of toxicity for the fetuses and 2-day-old neonates included the num ber of live, dead, and resorbed fetuses; body weights; gross abnorm alities; RBC, W BC, and blood cell differentials; Hgb levels; and the num ber of liver cells in the hem a topoietic differentiating, proliferating pool. The responses of 6-week-old progeny were evaluated on the basis of peripheral W BC and RBC counts, Hgb, and sm ears m a de from fem u r bone m a rrow or spleen to determ ine the num ber of cells in the differentiating, proliferating pool. Although not stated, the experim e ntal anim als m a y have been the sam e for both studies (Keller and Snyder, 1986, 1988), as all the details of the exposures were identical.

Benzene exposures m onitored hourly were well within target range (Keller and Snyder, 1988). There was no evidence of m a ternal toxicity at any exposure level. For the fetuses, there were no effects on litter size, m a le/fem ale ratio, or body weight or the num bers of dead, resorbed, or m a lform ed. Significant findings ( p <0.05) relevant to hem a topoietic toxicity in the three groups of progeny included the following.

Peripheral blood. Hgb levels of exposed and control m i ce were sim ilar 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 com p arable with those of the controls. The 2-day-old neonates

92exposed in utero to 5 ppm (16 m g /m 3 ) benzene exhibited significant increases in RBC counts and decreases in m ean corpuscular Hgb, and the 6-week-old adults exposed in utero to 5 and 20 ppm (64 m g /m 3 ) 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 num bers 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 m g /m 3 ) also had decreased num bers of late nucleated red cells and increased

num bers of nondividing granulocytes; these effects were not concentration dependent.

Hematopoietic organs. The 16-day fetuses from dam s exposed to benzene exhibited no changes in hem a topoietic param e ters of the liver at any benzene concentration. The 2-day-old neonates exposed in utero to 20 ppm (64 m g /m 3 ) benzene had changes in the liver that were significant at 20 ppm (64 m g /m 3 ). These included increase in the num ber of blast cells to 3.2 tim es the control value, depression in the num ber of late nucleated red cells (polychrom atic

norm oblasts and their nucleated progeny) to 70% of the control value, an increase in the num ber of lym phocytes to 1.4 tim es the control value, an increase in the num ber of nondividing granulocytes to 1.9 tim es the control value, and an increase in the num bers of dividing granulocytes to 2.5 tim es 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 m g /m 3 ) benzene also exhibited statistically significant changes in the cells of hem opoietic organs. These included fewer early nucleated red cells in the liver (to 55% of the control value), m o re splenic blast cells (to 6.5 tim es the control value), and increased dividing and nondividing splenic granulocytes (to 2.9 and 3.2 tim es the control value, respectively). The effects were not dose related and, with the exception of one cell count at 5 ppm (16 m g /m 3 ), were not statistically significant at lower concentrations.

The authors concluded that benzene induces hem a totoxicity in the offspring of m i ce exposed during pregnancy, as evidenced by reductions in the num bers of early nucleated red cells in the peripheral blood of 2-day-old neonates at all exposure levels and by the increase in the num ber 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 anim als, as clearly indicated by the near absence of such cells in the 6-weekold adult m i ce. A lim ited num ber of pregnant anim als (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) exam ined the influence of gender, developm ent, pregnancy, and ethanol consum ption on the hem a totoxicity of inhaled benzene. They exposed age-m a tched

m a le, virgin fem a le, and pregnant Swiss-W e bster m i ce (num ber of anim als in each group was not given, but there were 12 or 13 dam s) to 10 ppm (32 m g /m 3 ) for 6 hours/day for 10 consecutive days (GDs 6–15 for the pregnant fem a les). One-half of the anim als also received 5% ethanol in the drinking water during the exposure period. On day 11, bone m a rrow cells

93from the adults and liver cells from the fetuses were assayed for the num bers of CFU-E. CFU-E assays were also perform ed on bone m a rrow cells isolated from 6-week-old m a les and fem a les exposed in utero. Depressions in CFU-E num bers were seen only in m a les. CFU-E in adult

m a les 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 m a les exposed to benzene or benzene plus ethanol but not for ethanol alone. The CFU-E value for m a les exposed to both benzene and ethanol, however, was reduced by only 5% in com p arison with the control value as com p ared with 20% for ethanol alone. There were no significant differences for fem ale fetuses. Assays of CFU-E in 6-week-old

m i ce exposed to benzene in utero showed reductions of as m u ch as 50% in com p arison with controls for m a les, but there were no significant reductions for fem a les. In fact fem a les exposed to benzene and ethanol in utero showed a 40% elevation of CFU-E in com p arison with unexposed controls. For both ethanol and benzene, only a single concentration was used in these experim e nts. Thus, it is not possible to evaluate whether the results are part of a dose-response relationship.

Murray et al. (1979) conducted developm ental toxicity studies in m i ce that also included the evaluation of hem a tologic param e ters in the offspring. CF-1 m i ce were exposed to 500 ppm (1597 m g /m 3 ) benzene 7 hours/day on GDs 6–15 and were sacrificed on GD 18. There were no significant signs of m a ternal toxicity, and values for PCV, RBCs, Hgb, and W BCs were

com p arable with control values for the adult m i ce. There were no adverse effects on developm ental param e ters such as num ber of live fetuses or num ber of resorptions/litter; however, body weight of the m ouse 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 m ouse 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 m a lform ations. The investigators concluded that benzene did not induce

m a lform ations in m i ce at 500 ppm (1597 m g /m 3 ), 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 dem onstrate fetal toxicity at a dose that was not m a ternally toxic.

Ungvary and Tatrai (1985) exam ined the developm ental toxicity of benzene in CFLP

m i ce. Groups of 11 to 15 pregnant anim als inhaled 156 or 313 ppm (500 or 1000 m g /m 3 ) benzene 24 hours/day during GDs 6–15. Unexposed and air-exposed (cham b er) anim als served as controls. The m i ce were sacrificed on GD 18. The authors m e ntioned in the abstract of their paper that all solvents caused m oderate and concentration-dependent m a ternal toxicity. This was the only inform ation provided for m a ternal toxicity am ong the m i ce. Of the m ouse fetuses from dam s exposed to 156 and 313 ppm (498 and 1000 m g /m 3 ), 25 and 27%, respectively, exhibited weight retardation, and 10 and 11%, respectively, had retarded skeletal developm ent. All of these effects were statistically significant ( p <0.05). There were no differences between control and exposed groups in the incidence of m i nor anom alies or m a lform ations. The investigators concluded that under the conditions of this study, benzene does not induce m a lform ations, but it does induce fetal toxicity in m i ce exposed to 156 and 313 ppm (498 and 1000 m g /m 3 ), a

m a ternally toxic dose. The m ouse data suggest a LOAEL of 156 ppm (498 m g /m 3 ).

Murray et al. (1979) conducted developm ental toxicity studies in rabbits that also included the evaluation of hem a tologic param e ters in the offspring. New Zealand rabbits were

94exposed to 500 ppm (1597 m g /m 3 ) benzene 7 hours/day on GDs 6–18. The rabbits were sacrificed on GD 29. The rabbits exhibited no significant signs of m a ternal toxicity, and there were no adverse effects on developm ental param e ters such as num ber of live fetuses or num ber of resorptions/litter. Body weight and length of the rabbit fetuses were com p arable with control values. Two m i nor skeletal variants, lum b ar spurs and the proportion of fetuses with 13 ribs (the

norm al num ber is 12 or 13), occurred significantly less often am ong the exposed litters. Increases in the incidence of m a lform ations were not significant. Values for PCV, RBCs, Hgb, and W BCs were com p arable to control values for adult and fetal rabbits. The investigators concluded that benzene did not induce m a lform ations in rabbits at 500 ppm (1597 m g /m 3 ) 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 exam ined the developm ental toxicity of benzene and other solvents in New Zealand rabbits. Groups of 11–15 pregnant anim als inhaled 156 or 313
ppm (500 or 1000 m g /m 3 ) benzene 24 hours/day on GDs 7–20. Unexposed and air-exposed (cham b er) anim als served as controls. The rabbits were sacrificed on GD 30. The m a ternal rabbits exhibited concentration-related decreases in weight gain (to 37% of control values at 313
ppm [1000 m g /m 3 ]) and increases in relative liver weights (1.2 tim es control values at 313 ppm [1000 m g /m 3 ]); com p ared with control values, the effects were statistically significant at 313
ppm (1000 m g /m 3 ) ( p <0.05). Two dam s 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 m g /m 3 ]), concentration-related increases in the percent of dead or resorbed fetuses (3.1 tim es control values, p <0.05 at 313 ppm [1000 m g /m 3 ]), concentrationrelated increases in skeletal retardation (not statistically significant) and m i nor anom alies (2.5 tim es control values, p <0.05 at 313 ppm [1000 m g /m 3 ]), and concentration-related decreases in the percent of m a lform ations. The investigators concluded that under the conditions of this study, benzene does not cause m a lform ations, but it induces fetal toxicity in rabbits at 313 ppm (1000 m g /m 3 ), a m a ternally toxic concentration. The evidence supporting the LOAEL and the NOAEL in rabbits is weakened by a lack of experim e ntal details in the report, the sm all num bers of pregnant rabbits em ployed, and the use of the fetus rather than the litter as the experim e ntal unit.

4.2.2.3. Summary of Principal Reproductive/Developmental Toxicity Effects

As shown in the studies sum m a rized in the previous subsections and in Tables 7 and 8, adm i nistration of benzene has been associated with the onset of reproductive/developm ental effects in a wide range of experim e ntal anim als.

The susceptibility of reproductive organs to the toxic effects of benzene was shown in studies on the longer-term adm i nistration of benzene in which, am ong other organs, the rat testes (W olf et al., 1956) and the ovaries of fem a le CD-1 m i ce (W ard et al., 1985; NTP, 1986) were reduced in weight com p ared with controls and histopathologically altered as a result of benzene adm i nistration. The potential for these changes to affect reproductive function in m a les is suggested by reduced sperm count and increased num bers of abnorm al sperm when CD-1 m i ce inhaled up to 300 ppm (958 m g /m 3 ) benzene for 13 weeks.

95The experim e nts of Kuna et al. (1992) provided evidence that the inhalation of benzene by fem a le Sprague-Dawley rats before and during pregnancy resulted in reduced body and liver weights in weanlings, a valid reproductive/developm ental response because it occurred in the absence of any overt effects of the com pound on m a ternal toxicity. In addition, a considerable

num ber of experim e nts have addressed the potential of inhaled benzene to induce reproductive, developm ental, or teratogenic effects in the offspring of pregnant rodents exposed to the

com pound during the period of principal organogenesis. For exam ple, the studies by Kuna and Kapp (1981) of Sprague-Dawley rats point to the possibility of benzene-induced skeletal variations and abnorm alities in the fetuses of exposed dam s , although the quantitative extent of the effect was m a rginal at best. However, the weight of evidence for benzene as a developm ental toxicant arises from the extensiveness of the overall inform ation base. Several studies that em ployed essentially sim ilar experim e ntal protocols (see, e.g., Green et al., 1978) contribute to an overall picture of the com pound as inducing reduced fetal growth and delayed ossification. In addition, other studies reported reductions in fetal body weights, for exam ple, in Sprague-Dawley rats (Coate et al., 1984) and New Zealand white rabbits (Ungvary and Tatrai, 1985).

The developing hem a topoietic system in m i ce is affected by m a ternal exposure to benzene (Keller and Snyder, 1986, 1988; Corti and Snyder, 1996), which is consistent with benzene-induced hem a tologic abnorm alities in hum ans and adult anim als. The offspring of dam s exposed during gestation exhibited increased granulopoiesis and changes in the num bers of hem a topoietic progenitor and precursor cells, particularly during postnatal developm ent.

Multigenerational anim al studies on the reproductive/developm ental toxicity of benzene were not found in the literature, an inform ation gap that contributes to uncertainty about the overall context of the com pound’s reproductive and developm ental toxicity. As discussed in Sections 4.1.2.3 and 4.1.2.4, the evidence regarding reproductive and developm ental effects from hum an studies is lim ited. The hum an studies suffer from m a jor lim itations, including sm all

num bers of subjects, exposures to a m i xture of chem icals, and poorly defined benzene exposure levels and durations of exposure. In m a ny cases data on controls are lacking.

4.2.2.4. Mechanisms of Developmental and Reproductive Toxicity

The m echanism s for the developm ental and reproductive toxicity of benzene are not well understood. Sum m a rized below are a few of the suggested m echanism s that pertain specifically to the developm ental and reproductive toxicity of benzene.



### & Following adm i nistration of benzene to pregnant rats, Pushkina et al. (1968) observed decreased ascorbic acid content in the whole fetus and in m a ternal organs as the concentration of benzene increased, first in the m a ternal 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 m a ternal liver. The authors suggested that these alterations in ascorbic acid, RNA, and DNA content are possible m echanism s for fetal toxicity.



### & Ungvary and Donath (1984) suggested that dam a ge to the peripheral noradrenergic fibers observed in their study in pregnant rats m a y result in a disturbed control of ovarian and

96uterine blood flow and steroid production and m a y thus be instrum ental in the

em bryotoxic action of organic solvents.



### & Tatrai et al. (1980) suggested that several factors m a y be responsible for the

em bryotoxicity of benzene. First, because of its lipophilicity, benzene can pass the placental barrier and affect the em bryonal cells directly. Second, phenol, a m a jor

m e tabolite of benzene shown to inhibit DNA synthesis in bone m a rrow in vivo, can also pass the placental barrier. Third, benzene can dam a ge m a ternal circulation and cause bone m a rrow depression, resulting in adverse nutritional conditions for the fetus. Benzene oxide was suggested as the toxic m e tabolite of benzene, and it can be translocated from liver to blood and bone m a rrow. The enzym e s responsible for its production appear in the rat fetus late in pregnancy. However, in the hum an fetus, the
enzym e s are present during weeks 9–13 of pregnancy. The authors suggested that a transplacental effect is m o re plausible than any other m echanism

.



### & Reports indicating that paternal (as well as m a ternal) exposure to benzene is associated with increased risk for stillbirths and the findings of increased incidences of testicular lesions in benzene-exposed anim als suggest that exposure of m a les m a y be im portant in the reproductive toxicity of benzene (Savitz et al., 1989; Spano et al., 1989; W a rd et al., 1985).

4.2.3. Neurotoxicity

4.2.3.1. Oral Exposure

Sprague-Dawley rats given a single dose of 1870 m g /kg benzene exhibited trem ors and tonic-clonic convulsions and died within m i nutes. A dose of 352 m g /kg produced slight nervous system depression (Cornish and Ryan, 1965). The LD 50 for benzene in nonfasting rats is 0.81 g/kg.

Hsieh et al. (1988a) evaluated the effects of benzene on neurotransm itters of the brain. CD-1 adult m a le m i ce (five/group) received benzene in the drinking water for 4 weeks ad lib. Controls received tap water. The nom inal concentrations were 0, 40, 200, and 1000 m g /L. However, based on water consum ption, the actual daily estim ated doses of benzene were 0, 8, 40, and 180 m g /kg/day.

The benzene concentrations used in this study did not significantly alter behavior, body weights, or food and water consum ption. Generally, the increases in levels of m onoam i ne neurotransm itters were dose related at 8 and 40 m g /kg/day, but in several cases there were no further increases at 180 m g /kg/day. The increases were greatest for norepinephrine in the hypothalam u s (increased over control by approxim a tely 38, 55, and 58% at 8, 40, and 180

m g /kg/day, respectively) and m e dulla oblongata (16, 42, and 20%). For serotonin, a sim ilar pattern of increase was observed in the higher association center with increased exposure to benzene. In the hypothalam u s, the serotonin levels were 21, 86, and 93% above the control values. A generalized increase was also observed in the m e dulla oblongata (5, 25 and 19%) and the m i dbrain (8, 46, and 23%). The increases in the parent com pounds were associated with increases in their corresponding m e tabolites, reflecting increased turnover of the am ines. The

97data from this study did not establish a no-effect level of benzene in drinking water. In fact, although the assay has potential as a biom arker for exposure, the biological significance of these findings is questionable. The findings, however, suggest that neurotoxicological effects in the hypothalam u s and the peripheral nervous system could be of particular concern for the developing central nervous system , resulting in perm anent effects.

4.2.3.2. Inhalation Exposure

Ten rabbits exposed to ~ 45,000 ppm (143,760 m g /m 3 ) benzene exhibited light narcosis after 3.7 m i nutes of exposure, followed by trem ors, chewing, excitem e nt, and running

m ovem e nts after 5 m i nutes (Carpenter et al., 1944). There was a loss of pupillary reflex to strong light after 6.5 m i nutes, loss of blink reflex to tactile stim ulus after 11.4 m i nutes, pupillary contraction after 12 m i nutes, involuntary blinking after 15.6 m i nutes, and death after 36.2

m i nutes (Carpenter et al., 1944).

Andersson et al. (1983) exam ined the effects of high concentrations of benzene on dopam i ne and noradrenaline turnover within various parts of the hypothalam u s that are involved in neuroendocrine regulation. Turnover of the catecholam ines was assessed by m easuring changes in the degree of catecholam ine stores following tyrosine hydroxylase inhibition. Sprague-Dawley rats exposed to 1500 ppm (4792 m g /m 3 ) benzene 6 hours/day for 3 days were divided into two groups of four or six anim als. In the first group (–inhibitor), rats that served as controls were sacrificed im m e diately following exposure. In the second group (+inhibitor), the injection of tyrosine hydroxylase inhibitor im m e diately followed exposure, and the anim als were sacrificed 2 hours later. The extent of depletion of the catecholam ines was determ ined by calculating the percentage depletion in the test group, based on the levels of catecholam ine present in the control group at the tim e of the injection of the hydroxylase inhibitor.

Quantitative m i crofluorim e try dem onstrated that benzene induced statistically significant alterations in catecholam ine content and turnover in various sections of the hypothalam u s. Benzene (–inhibitor) produced increases in the catecholam ine fluorescence in the m e dian and lateral palisade zones of the m e dian em inence ( p <0.05) and within the posterior periventricular hypothalam i c region ( p <0.05). Benzene (+inhibitor) enhanced the disappearance of catecholam ine in the m e dian palisade zone ( p <0.002), within the posterior periventricular hypothalam i c region ( p <0.05), within the parvocellular part of the paraventricular hypothalam i c nucleus ( p <0.01), and within the dorsom edial hypothalam i c nucleus ( p <0.01).

The investigators concluded that benzene produced a pattern of discrete changes in noradrenalin and dopam i ne turnover in certain areas of the hypothalam u s. The study included only one concentration of benzene and a sm all num ber of anim als. No LOAEL or NOAEL was established. Tyrosine hydroxylase is the key enzym e for biosynthesis of catecholam ine, and the hypothalam u s is one of the m a jor association centers in the central nervous system

. It is clear that benzene will affect the functions of all th ese centers at an inhalation dose of 500 ppm (1597
mg / m 3 ), as dem onstrated in the changes in dopam i ne 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 m g /m 3 ) benzene 8

98hours/day on GDs 8–10 ppm

. The abundance of fluorescent noradrenergic fibers norm ally 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 cham ber of the eye. The benzene-induced dam a ge to the sym p athetic nerve-plexus of the iris was sim ilar to that in the uterus and ovary. The density of fluorescent fibers was decreased (dose related), and the iris exhibited substantial hyperem i a 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 em bryotoxicity.

Dem p ster et al. (1984) exam ined the tem poral relationship between the behavioral and hem a tologic effects of inhaled benzene. Mice (30–45/group) were exposed to benzene concentrations of 100–3000 ppm (319–9584 m g /m 3 ) 6 hours/day for the num ber of days necessary to achieve a concentration × tim e product of 3000 ppm -days. The controls were agem a tched and exposed to air. The behavioral param e ters evaluated included m ilk licking (an observation of the m i ce licking a spout protruding from a wall in order to obtain m ilk), hindlim b grip strength (a m ouse 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 hom e cage food and water intake (m easured by weighing water bottles and feeders daily). Lym phocyte num bers were reduced to 68% of control values after five exposures to 100 ppm (319 m g /m 3 ) benzene, to 50% after two exposures to 300 ppm (958 m g /m 3 ), and to 50% after one exposure to 1000 and 3000 ppm (3194 and 9584 m g /m 3 ). Maxim a l depression in the lym phocyte counts occurred after 10 days of exposure to 300 ppm (958 m g /m 3 ). The lym phocyte counts

rem a ined depressed throughout each exposure regim e n.

RBC counts were not depressed as rapidly or to the sam e extent as were the lym phocytes. At the end of exposure to 300 ppm (958 m g /m 3 ), for exam ple, the RBC counts were reduced to 70% of control values. Increased m ilk licking, the m o st sensitive of the behavioral param e ters, was statistically significant following 1 or 2 days of exposure to 100 ppm (319 m g /m 3 ) and after 4 or 5 days of exposure to 300 ppm (958 m g /m 3 ). The m a xim a l increase in this behavior pattern occurred after 7–8 days of exposure to 300 ppm (958 m g /m 3 ), following the sam e tim e course as the hem a tologic effects. Short-term exposures to high concentrations did not increase m ilk licking, but they did increase food consum ption. Thus the increased m ilk licking at 100 ppm (319 m g /m 3 ) was apparently not due to hunger. Exposure to 1000 ppm (3195 m g /m 3 ) for one exposure reduced hindlim b grip strength. These effects disappeared following term ination of
exposure.

In a study designed to reflect occupational exposure, m a le CD-1 and C57BL/6 m i ce were exposed to 300 or 900 ppm (958 or 2875 m g /m 3 ) benzene 6 hours/day for 5 days followed by 2 weeks of no exposure, after which the exposure regim e n was repeated for an unspecified am ount of tim e (Evans et al., 1981). Seven categories of behavioral activities were m onitored in exposed and control anim als: stereotypic behavior, sleeping, resting, groom ing, eating, locom o tion, and fighting. Only m i nim a l and insignificant differences were observed between the two strains of

m i ce. Increased behavioral activity in both strains was observed after exposure to benzene. Mice exposed to 300 ppm (958 m g /m 3 ) benzene had a greater increase than those exposed to 900
ppm (2875 m g /m 3 ), probably because of narcosis-like effects induced at the higher exposure

99level. It is not known whether benzene induces behavioral changes by direct action on the central nervous system

. Frantik et al. (1994) found that m a le albino SPF rats from a W i star-derived strain exposed to benzene for 4 hours in glass cham bers (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 m g /m 3 ). W h en fem a le H strain m i ce were exposed to benzene for 2 hours, the 30% effect level for depression of evoked electrical activity in the brain was 856 ppm (2735 m g /m 3 ).

Adult m a le Kunm ing m i ce (five/group) were exposed to 0, 0.78, 3.13, or 12.52 ppm (0, 2.5, 10, 40 m g /m 3 ) benzene 2 hours/day, 6 days/week for 30 days (Li et al., 1992). Exposures were conducted under static conditions in 300 m 3 plexiglass cham bers. Benzene was m onitored by gas chrom atography every 30 m i nutes for 3 days at the beginning of the experim e nt, but apparently it was not m onitored for the rest of the experim e nt. The anim als were m onitored for Y m aze perform ance (rapid response); locom o tor activity; forelim b grip strength; acetyl cholinesterase (AChE) activity in the blood and brain; brain, liver, spleen, and kidney weights; and bone m a rrow cellularity. Statistical significance was evaluated by t value for forelim b grip strength and locom o tor activity and by U value for rapid response frequency.

Significantly increased grip strength was observed at 0.78 ppm (2.5 m g /m 3 ), but at higher doses, grip strength was significantly decreased. The frequency of rapid response (running a Y

m aze in less than 3 seconds) followed a pattern sim ilar to forelim b 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 m g /m 3 ) groups, respectively. No statistically significant differences were observed in AChE activity in either blood or brain or in locom o tor 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 m a rrow histologic investigation was presented. There were no apparent responses in the 0.78 or 3.13 ppm (2.5 or 10 m g /m 3 ) groups. In the high-dose 12.52 ppm (40 m g /m 3 ) group, however, there were reductions of 91% in m y eloblasts, 64% in
prem y eloblasts, 77% in m e tam y elocytes, 100% in reticulum , and 100% in erythroblasts.

The neurological effects reported in this study—forelim b grip strength and frequency of rapid response—are unique and interesting. However, several lim itations in this study prevent its use in establishing an RfC. The num ber of anim als used in each group (five) was low, exposures were perform ed under static conditions, and benzene concentrations were m onitored on only the first 3 of 30 exposure days. The very large decreases in several blood param e ters are in contrast with the findings of m o st other studies, which found m i nim a l or no response in bone

m a rrow param e ters at sim ilar exposure concentrations. Most other studies on the hem a totoxicity of inhaled benzene used exposure durations of 6 hours/day, in contrast to the 2 hours/day in this study. Thus, the dram a tic effects on bone m a rrow param e ters in this study suggest that actual benzene exposure m a y have been higher than reported.

1004.2.3.3. Summary of Neurotoxic Effects

The database for establishing threshold dose levels for neurotoxicity effects is lim ited. Neurological effects of benzene have been observed in anim als and hum ans, and no clear NOAEL could be identified in any species. No com p lete neurological testing has been conducted in anim als or hum ans. The exposure levels used in the neurotoxicity studies

sum m a rized 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 com p laints in workers exposed to 6–16 ppm (20–52m g/m 3 ), but no objective evaluation was conducted, exposure was poorly characterized, and there was confounding exposure to low levels of toluene (1.6 ppm [5 m g /m 3 ]). Other hum an neurotoxicity studies reported m u ch higher exposure levels. No com p lete functional observational battery has been conducted in anim als. Dem p ster et al. (1984) reported reduced

m ilk licking in m i ce exposed to 100 ppm (319 m g /m 3 for 1 or 2 days, but grip strength was not affected until exposure reached 1000 ppm (3195 m g /m 3 ). Li et al. (1992) reported effects on grip strength, locom o tor activity, and Y-m aze perform ance (rapid response) at 0.78 ppm (2.5 m g /m 3 ), but poor exposure characterization and m a rked inconsistencies between this and other studies’ hem atologic effects m ean that this value is unreliable.

4.2.4. Immunotoxicity

4.2.4.1. Oral Exposure

Male Charles River CD-1 m i ce (five/group, 6–7 weeks of age) were exposed to 0, 31, 166, or 790 m g /L (0, 8, 40, or 180 m g /kg/day) benzene in drinking water for 28 days (Hsieh et al., 1988b). The treatm e nt had no adverse effects with respect to m o rtality, 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 thym us weights were decreased at all exposure levels, the values were not statistically significantly different from control values. Dose-related hem a tologic effects (erythrocytopenia, leukocytopenia, lym phocytopenia, increased MCV) were observed at all exposure levels. The authors indicated that the increased MCV and decreased HCT and num bers of RBCs were indicative of severe m acrocytic anem ia.

Biphasic responses were observed in im m unological tests, including m itogen-stim ulated (lipopolysaccharide [LPS], pokeweed m itogen, concanavalin A [Con A], phytohem a glutinin [PHA]) splenic lym phocyte proliferation; m i xed splenic lym phocyte culture response to allogenic yeast artificial chrom osom e [YAC]-1 cells; cytotoxic splenic T lym phocyte response to allogenic YAC-1 cells with a significantly increased response at the low-exposure level; and significantly decreased responses at the m i d- and/or high-exposure level. Using several m e thods to determ ine prim a ry antibody response to SRBC, significantly decreased responsiveness was observed at the m i d- and/or high-exposure levels. This response was either significantly increased or not different from controls in m i ce at the low-exposure level. This study identified a LOAEL of 8 m g /kg/day (the lowest dose tested) for hem a tologic and im m unological effects in

m a le m i ce exposed to benzene in drinking water for 30 days. No NOAEL was established.

101In a subchronic study, groups of m a le C57BL/6 m i ce were exposed via drinking water to 0, 152, or 853 m g /L benzene for 7–28 days. Using estim ated daily water intakes, the authors calculated benzene dosages of 27 and 154 m g /kg/day (Fan, 1992). Five m i ce per group were sacrificed after 7, 14, 21, or 28 days of exposure. An unspecified num ber of m i ce were exposed to 152 m g /L for 28 days and sacrificed 7, 14, or 21 days after the last dosage. The focus of this study was to determ ine the toxicity of benzene on natural killer (NK) cells involved in nonspecific host resistance and on IL-2, which is the prim a ry 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 m i ce. Significant decreases in the num ber of spleen cells were observed in both groups of benzene-exposed m i ce. This effect was observed after 21 days of exposure in the 152 m g /kg/day group and after 14 days in the 852 m g /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

num bers and IL-2 production were also depressed in the m i ce exposed to 152 m g /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 m g /kg/day (the lowest dose tested) for effects on the im m une system in m a le m i ce. A NOAEL was not identified.

Fem a le B6C3F1 m i ce (12/group, 6–7 weeks of age) were exposed to benzene in drinking water (containing em ulphor to increase solubility of benzene) at levels of 0, 50, 1000, or 2000

m g /L (0, 12, 195, or 350 m g /kg/day, as calculated by the authors) for 30 days (W hite 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 m i d- and high-exposure levels. Dose-related leukopenia and lym phocytopenia 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 glutam ic oxaloacetic transam i nase, or serum glutam ic pyruvic
transam i nase, indicators of renal and hepatic dam a ge.

Dose-related changes were observed in im m unological tests on spleen cells and in assays of bone m a rrow ( p <0.05); decreases were observed with respect to IgM antibody form ing cells/spleen in response to SRBC, lym phocyte proliferation response to the T cell m itogen Con A and the B cell m itogen LPS, num ber of T lym phocytes, and fem o ral GM-CFU; and an increase was observed in bone m a rrow cell DNA synthesis. These effects were not significant at 12

m g /kg/day but were dose related ( p <0.05). Of all the im m unological indices tested, only one endpoint (stim ulation index for lym phocyte proliferation of spleen cells in response to m e dium containing 0.5 

###  g/m L Con A) was significantly decreased at 12 m g /kg/day ( p <0.05). The

num ber of B lym phocytes was not affected, but the investigators com m e nted that the num ber of B lym phocytes in the controls was lower than for historical controls for their laboratory. This study identifies a NOAEL of 12 m g /kg/day and a LOAEL of 195 m g /kg/day for hem a tologic effects in m i ce exposed to benzene in drinking water for 30 days, and a LOAEL of 12 m g /kg/day for im m unological effects.

1024.2.4.2. Inhalation Exposure

Male C57BL/6J m i ce (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 m g /m 3 ) in whole-body dynam i c inhalation cham bers 6 hours/day for 6 days (Rozen et al., 1984). Mice were bled within 30–90 m i nutes of the last exposure to determ ine peripheral blood counts. Five anim als with blood counts closest to the group m ean were selected for lym phocyte assays. Single-cell suspensions of bone m a rrow and spleen cells from these anim als were used for T and B cell enum eration and m itogeninduced proliferative assays. Statistical significance ( p <0.05) was determ ined by one-way ANOVA.

Lym phocyte counts were depressed at all exposure levels. Lym phocyte counts were significantly reduced to approxim a tely 65% of controls in the 10.2 ppm (32.6 m g /m 3 ) exposure group and to 35% of controls at 100 or 300 ppm (319 and 962 m g /m 3 ). Erythrocyte counts were significantly stim ulated (115% of control levels) at 10.2 ppm (32.6 m g /m 3 ) and were depressed only at 100 and 300 ppm (319 and 962 m g /m 3 ). At exposures of 10.2 ppm (32.6 m g /m 3 ), the frequency of fem o ral B lym phocyte colony-form ing cells was reduced to approxim a tely 30% of the control value. In contrast, the num ber of fem o ral B lym phocytes was not significantly reduced in the low-exposure group, but the num ber was reduced to less than 10% of the control value at 100 or 300 ppm (319 and 962 m g /m 3 ). Sim ilarly, splenic PHA-induced blastogenesis was significantly depressed at 31 ppm (99 m g /m 3 ) without a concom itant depression in num bers of T lym phocytes. The num bers of T lym phocytes were reduced to less than 50% of controls in the 100 and 300 ppm (319 and 962 m g /m 3 ) exposure groups. These results dem onstrate that short-term exposure to inhaled benzene even at low exposure concentrations can cause reductions in im m une-associated processes. This study identifies a LOAEL for depression of lym phocytes of 10.2 ppm (32.6 m g /m 3 ), the lowest dose tested.

Rosenthal and Snyder (1985) investigated the effects of exposure to benzene at 0, 10, 30, 100 (320 m g /m 3 ) or 300 ppm (960 m g /m 3 ) on the im m une response of m a le C57BL/6J m i ce (5–7/group) to challenge with the facultative intracellular pathogen Listeria monocytogenes

. All

m i ce (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 m easured at 1, 4, and 7 days after infection as an index of the resistance of the m i ce to infection. Body and spleen weights were determ ined, T and B lym phocytes were enum erated, spleen cellularity was determ ined, and spleen

m onocyte/m acrophage, polym orphs, lym phocytes, and nucleated RBCs were scored.

None of the benzene exposure treatm e nts 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 / m 3 ) or m o re 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
treatm e nts. The authors suggested that benzene exposure caused a delay in the cell-m e diated im m une response, as there was a tem porary increase in spleen bacterial counts. Both T and B

103lym phocytes were particularly sensitive to benzene exposure. Lym phocyte 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 m g /m 3 ) ppm and a NOAEL of 10 (32 m g /m 3 ) ppm for effects on the im m une system

.

4.2.4.3. Summary of Immunotoxic Effects

As detailed above, dose-related adverse effects on spleen weight and cellularity and on various m easures of im m une 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; W h ite et al., 1984). Dose-related adverse effects on several m easures of im m une function have been observed, including lym phocyte count, m itogen-stim ulated lym phocyte proliferation, prim a ry antibody response to SRBCs, IL-2 production, fem o ral B lym phocyte colony-form ing cells, num ber of T lym phocytes, and proliferation of L. monocytogenes bacteria in the spleen (Hsieh et al., 1988b; Fan, 1992; W h ite et al., 1984; Rozen et al., 1984). The results indicate that exposure to benzene, whether oral or inhaled, adversely affects the im m une 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 adm i nistration. Hem a totoxicity and im m unotoxicity have been consistently reported to be the m o st sensitive endpoints of noncancer toxicity in lim ited studies in hum ans and experim e ntal anim als, and these effects have been the subject of several reviews (Aksoy, 1989; Goldstein, 1988; Snyder et al., 1993b; Ross, 1996). The bone m a rrow is the target organ for the expression of benzene hem a totoxicity and im m unotoxicity.

Chronic exposure to benzene results in progressive deterioration in hem a topoietic function. Anem ia, leukopenia, lym phocytopenia, throm bocytopenia, pancytopenia, and aplastic

anem ia have been reported after chronic benzene exposure (see detailed discussions and

sum m a ries 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 m a rrow hyperplasia. AML has been frequently observed in studies of hum an cohorts exposed to benzene, and there is evidence linking benzene exposure to several other form s of leukem i a (U.S. EPA, 1998a). W h ether the hem a totoxic/im m unotoxic effects of benzene and its carcinogenic effects are caused by a com m on m echanism is not yet known, due in part to the fact that although the bone m a rrow depressive effects of benzene in hum ans can be readily duplicated in several experim e ntal anim al

m odel system s, a suitable experim e ntal anim al system for the induction of leukem i a has not yet been developed. In addition, the hem a totoxicity/im m unotoxicity of benzene leads to significant health effects apart from potential induction of leukem i a.

Although the decreased ALC and leukem i as observed from benzene exposure both result from bone m a rrow toxicity, they do not necessarily result from the sam e m echanism s , and decreased ALC m a y not be a necessary precursor for leukem i a. For exam ple, decreased ALC

m a y be the result of cytotoxicity independent of the interaction of benzene m e tabolites with DNA and/or DNA-associated proteins and tum o r form ation. Thus, integration of the cancer and

104noncancer health assessm ents of benzene is not considered possible given the current state of knowledge regarding m ode of action.

Leukocytopenia has been consistently shown to be a m o re sensitive indicator of benzene toxicity than anem ia in experim e ntal anim al system s, and lym phocytopenia has been shown to be an even m o re sensitive indicator of benzene toxicity than is overall leukocytopenia (Snyder et al., 1980; W a rd et al., 1985; Baarson et al., 1984)

. Rothm an et al. (1996a) also found that a decrease in absolute lym phocyte count was the m o st sensitive indicator of benzene exposure in a group of workers. Sim ilarly, W a rd et al. (1996) dem onstrated a strong relationship between benzene exposure and decreased W BC counts in a rubber worker cohort, but no significant relationship with RBC counts was found. A com m on observation in experim e ntal anim al and hum an studies on benzene and lym phocytopenia—except for the occupational study of Collins et al. (1991)—has been the absence of a clear threshold.

Hum a n and experim e ntal anim al studies indicate that benzene affects both hum oral and cellular im m unity. Dose-related adverse effects on spleen weight and cellularity and various
param e ters of im m une function have been observed following both oral and inhalation exposure to benzene.

Benzene toxicity to hum ans exposed in the workplace has been characterized as having either early hem a totoxicity or, with prolonged exposure to high doses, irreversible bone m a rrow dam a ge. Studies of worker populations in factories in which benzene was used as a solvent (Snyder and Kocsis, 1975) showed a range of hem a totoxic effects, including anem ia, leukopenia, and throm bocytopenia. In som e cases, m o re than one cell type was decreased. A decrease in the levels of all the classes of blood cell types in circulation is term ed pancytopenia and is usually associated with irreversible bone m a rrow aplasia. Aplastic anem ia is fatal in m o st cases. In those who survive aplasia, the m a rrow appears to be dysplastic. Myelodysplastic syndrom e, which has been called preleukem i a, is likely an early stage of AML.

The evidence is strong that benzene m e tabolism plays a critical role in toxicity (Snyder and Hedli, 1996). Hepatic m e tabolism plays an im portant role in toxicity (Andrews et al., 1977). In addition to hepatic m e tabolism
, it appears that secondary m e tabolism of benzene m e tabolites in bone m a rrow contributes to toxicity (Irons et al., 1980; Schlosser and Kalf, 1989). Thus, elucidation of the m e tabolic pathway for benzene biotransform ation is essential for a full understanding of the m echanism of toxicity.

There are alternative routes by which the first step of benzene m e tabolism
, nam e ly, phenol form ation, can occur. CYP2E1, and perhaps other cytochrom es P450, can generate H 2 O 2 when acting as oxidases of nicotinam ide adenine dinucleotide phosphate (NADPH). The hydroxyl radical form ed from H 2 O 2 can hydroxylate benzene to yield phenol. An alternative

m echanism for phenol form ation reflects on the fate of the benzene oxide-oxepin system (see Figure 1). W h en benzene oxide is the first product, it can rearrange nonenzym atically to form phenol. Alternatively, benzene oxide can be further m e tabolized 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 form ation of the prem e rcapturic acid. It is likely that benzene oxide or its oxepin (interm e diate) are precursors to ring opening (W itz et al., 1996). Phenol can be further hydroxylated to form

105hydroquinone or catechol. In theory, 1,2,4-benzenetriol m a y be form ed 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 hum ans.

Any of the phenolic m e tabolites m a y be conjugated with either sulfate or glucuronide. In addition to L-phenylm e rcapturic acid reported by Parke and W illiam s (1953), other

m e rcapturates include 6-N-acetylcysteinyl-S-2,3-cyclohexadienol (Sabourin et al., 1988a, b) and 2,5-dihydroxy-phenyl-m ercapturic acid (Nerland and Pierce, 1990). The urine also contains two ring-opening products, MA (Drum m ond 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., N 7 -phenylguanine (Norpoth et al., 1988).

The production of benzene m e tabolites, largely in the liver, is followed by their transport to the bone m a rrow and other organs. There are m a ny possible causes of bone m a rrow toxicity. Pfeifer and Irons (1983) suggested that covalent binding of hydroquinone to spindle fiber protein could explain inhibition of cell replication by benzene. Dam a ge to DNA could result in bone

m a rrow depression, leading to aplastic anem ia, which in survivors leads to m a rrow dysplasia and ultim ately to AML (Snyder and Kalf, 1994). Benzene m e tabolites could cause dam a ge to DNA by two possible m echanism s. One pathway focuses on the m e tabolic activation of benzene to species that covalently bind to DNA to produce m u tagenic events that are expressed as leukem i a. The second m echanism involves the production of m e tabolites that cause oxidative stress, subsequent oxidative dam a ge to DNA, and a m u tagenic effect that has the sam e consequences.

W ithin the bone m a rrow, both hem a topoietic progenitor cells and bone m a rrow strom al cells are potential targets of benzene toxicity (Aksoy, 1988; Gaido and W i erda, 1985). Progenitor cells are thought to be the cells of origin for leukem i as (Greaves, 1993) and are attractive as potential targets of toxins such as benzene, whose toxic effects are not restricted to a single hem a topoietic lineage. Strom al cells are critical in the regulation of norm al hem opoiesis (Dorshkind, 1990) and have been considered as im portant targets of benzene toxicity (Gaido and W i erda, 1985). Strom al cells are intim ately associated with developing blood cells and regulate hem opoiesis via direct cell-to-cell interactions, the production of extracellular m atrix

com ponents, and the secretion of soluble m e diators such as cytokines and eicosanoids (Billips et al., 1991).

Although m e tabolism is central to benzene toxicity, studies of the m e tabolic capability of hum an bone m a rrow are scarce, particularly in hum an strom al and CD34 + cells.

Benzene hem a totoxicity is a com p lex process that m o st likely involves interaction am ong several m e tabolites (Billips et al., 1991). Possible interm ediates considered at present to be im portant in benzene hem a totoxicity are polyhydroxylated benzene m e tabolites, for exam ple, hydroquinone and 1,2,4-benzenetriol; their quinone oxidation products, for exam ple, p

benzoquinone, form ed via oxidation of hydroquinone, and sem i quinone free radical interm ediates form ed during the oxidation of polyhydroxylated m e tabolites to quinones; reactive oxygen species form ed during the oxidation of the polyhydroxylated m e tabolites; and ringopened aldehydic benzene m e tabolites, for exam ple, MUC and 6-hydroxy- trans, trans -2,4- hexadienal.

106MUC and its m e tabolite 6-hydroxy- trans, trans -2,4-hexadienal are reactive ring-opened hem a totoxic com pounds. They exhibit a host of biological activities that could potentially be im portant in their m echanism s of toxicity in relation to benzene.

Benzene hem a totoxicity occurs when its hepatic m e tabolites (Sam m e tt et al., 1979)—phenol, catechol, and hydroquinone—are transported to the bone m a rrow (Greenlee et al., 1981) and further oxidized in a peroxidase-m ediated (Sm ith et al., 1989) reaction to biologically reactive interm ediates that can potentially affect hem a topoiesis.

The ability to alter cytokine-dependent growth and differentiation in hem a topoietic progenitor cells appears to be a property of m a ny agents, with leukem ogenic potential for hum ans (Irons and Stillm an, 1993). Several studies have reported significant effects of benzene on hem a topoietic stem and progenitor cells (Cronkite et al., 1989). In one study (Seidel et al., 1989), a dose-dependent depression of all stem cell com p artm ents was observed in BDF1 m i ce exposed for 16 weeks to airborne concentrations of benzene as high as 99 ppm (317 m g /m 3 ) for 6 hours/day, 5 days/week. The results of this study indicated that the GM-CFU was m u ch less sensitive than the erythroid CFUs at higher doses of benzene.

Another study (Dem pster and Snyder, 1990) reported that short-term exposure of m i ce to benzene induced a shift toward granulocytic differentiation, a growth advantage for granulocytic progenitor cells in the bone m a rrow and spleen, and to a resultant increase in the total num ber of granulocytes. These results suggest that benzene, or hydroquinone, is acting on the m y eloid
stem or progenitor cells. The adm i nistration of benzene, or hydroquinone, specifically stim ulates granulopoiesis in m i ce and induces granulocytic differentiation in m y eloblasts of the hum an prom yelocytic leukem i a cell line, HL-60, as well as the norm al m u rine IL-3 dependent

m y eloblastic cell line, 32D.3 (G). Benzene and hydroquinone do this by replacing the requirem e nt for granylocyte colony-stim ulating factor and leukotriene D 4 , respectively, for induction of differentiation.

In conclusion, benzene-induced m echanism s of hem a totoxicity are poorly understood. Most proposed m echanism s of benzene-induced effects such as cytotoxicity, apoptosis,

m

u tagenesis, and cell replication can be dem onstrated in vitro at relatively high concentrations of benzene m e tabolities (Snyder et al., 1993b). Benzene effects in vitro do not necessarily correlate with benzene-induced hem a totoxicity following in vivo exposures to low concentrations, as in vitro assays of hem a topoiesis lack the com p lex hem a topoietic cell interrelationships, fine-tuned regulation, and com p ensation m echanism s that are essential in norm al hem a topoiesis. The m o st im portant aspects of benzene-induced hem a totoxicity are the possible toxic effects caused by low concentrations of benzene. Concurrent assessm ent of toxicity elicited by high concentrations of benzene (100 [320 m g /m 3 ] and 200 [640 m g/m 3 ] ppm ) established a reference for com p arison.

In com m on with m a ny other organic solvents, benzene has been shown to produce neurotoxic effects in test anim als and hum ans after short-term exposures to relatively high concentrations. The neurotoxicity of benzene, however, has not been extensively studied, and no system atic studies of the neurotoxic effects of long-term exposure were located. Benzene produces generalized sym p tom s such as dizziness, headache, and vertigo at levels of 250–3000
ppm (799–9584 m g /m 3 ) (Brief et al., 1980), leading to drowsiness, trem or, delirium
, and loss of

107consciousness at 700–3000 ppm (2236–9584 m g /m 3 ). These neurological sym p tom s are reversible upon rem oval 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 m g /m 3 ) com p lained of frequent headaches, becam e tired easily, had difficulties sleeping, and com p lained of m e m o ry loss. The lim itations 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 anim als have exam ined the neurotoxicity of benzene, and those that have used short-term exposures, usually to concentrations that have been shown to induce significant hem a totoxicity. One exception is the study by Li et al. (1992), which observed biphasic responses in forelim b grip strength and frequency of rapid response in running a Y

m aze following inhalation exposure to 0, 0.78, 3.13, or 12.52 ppm (0, 2.5, 10, or 40 m g /m 3 ) benzene 2 hours/day, 6 days per week for 30 days. Both responses increased at the lower concentration and declined at the interm ediate and high concentrations. There were, however, a

num ber of uncertainties in the experim e ntal protocol that need to be resolved before these observations can be used in establishing NOAEL or LOAEL values.

No com p lete neurological testing has been conducted in anim als and hum ans. 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 som e evidence of reproductive and developm ental benzene toxicity from hum an epidem iology studies, but the data did not provide conclusive evidence of a link between exposure and effects. In m o st cases, there was exposure to other chem icals as well, and the quantitative data were not sufficient to determ ine a NOAEL or a LOAEL. Som e test anim al studies provide lim ited evidence that benzene affects reproductive organs; however, these effects were lim ited to high-exposure concentrations that exceeded the m a xim u m tolerated dose. Fertility studies that have shown adverse effects on the num ber of live fetuses used benzene concentrations that caused severe m a ternal 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 anim als are fairly consistent across species and dem onstrate that benzene is fetotoxic and causes decreased fetal weight and/or m i nor skeletal variants at concentrations of greater than 47 ppm (150 m g /m 3 ) (Coate et al., 1984; Green et al., 1978; Kuna and Kapp, 1981; Murray et al., 1979). Exposure of m i ce to benzene in utero during developm ent has been shown to cause changes in the hem a togenic 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 experim e ntal design lim itations (see detailed discussion of studies in Section 4.2.2).

Although benzene exposure has been shown to result in structural and num erical chrom osom al abberations in hum an lym phocytes, the quantitative relationship between

m easured benzene exposures and clastogenic effects in hum ans is unknown. Associations between benzene exposure and genotoxic effects in hum ans under occupational conditions have been dem onstrated. In anim al studies benzene has been shown to induce cytogenic effects, including chrom osom e and chrom atid aberrations, sister chrom atid exchanges, and m i cronuclei

108in vivo and in vitro. Mutagenicity of benzene m e tabolites is well established. However, benzene itself is not m u tagenic in bacterial and anim al system s (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 inform ation regarding whether infants and children m a y be m o re susceptible than adults to the hem a totoxic effects of benzene. Developm ental toxicity was described in Section 4.2.2. For exam ple, inhalation exposure of benzene in pregnant rats on GDs 6–15 resulted in fetotoxicity at 100 ppm (319 m g /m 3 ). The study established a LOAEL of 100
ppm (319 m g /m 3 ) and a NOAEL of 40 ppm (128 m g /m 3 ) for the fetal toxicity of benzene. Benzene could be a potential risk factor for the developm ent of childhood leukem i a (see reviews, OEHHA, 1997; Sm ith and Zhang, 1998; U.S. EPA, 1998a). Evidence from hum an and anim al studies (McKinney et al., 1991; Shaw et al., 1984; Shu et al., 1988; Buckley et al., 1989) suggests that increases in childhood leukem i a m a y be associated with in utero exposures and

m a ternal and paternal exposure prior to conception.

Evidence in anim als suggests that exposure to benzene in utero alters m a turation of lym phocytes, erythrocytes, and granulocytes (OEHHA, 1997). The consequences of in utero exposure to benzene can be detected as alterations in cell population num bers and functional properties into adulthood. Dam a ge during the initial in utero stages of hem a topoiesis could have lasting effects, as has been dem onstrated for a num ber of other toxicants (OEHHA, 1997).

Exposure of m i ce to benzene in utero during developm ent has been shown to cause changes in hem a togenic 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 dam a ge to the hem a topoietic system can occur during developm ent. It is not known whether the effect is the sam e as in adults or is unique to developm ental exposures. Dose-related hem a totoxicity was observed in the absence of any apparent m a ternal effects in Swiss-W e bster

m i ce (Keller and Snyder, 1986, 1988). At high doses, significantly increased skeletal variations in CF-1 m ouse 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 inform ation on peripheral blood effects. The absence of a clear NOAEL for fetal/neonatal exposure m a kes it difficult to directly com p are fetal/neonatal and adult toxicity.

There are few data on the effects of direct exposure of children to benzene. However,

som e indirect evidence suggests that children m a y be susceptible to benzene-induced hem a totoxicity. There is m ounting evidence that key changes related to the developm ent of childhood leukem i a occur in the developing fetus. Several studies have reported that genetic changes related to eventual leukem i a developm ent occur before birth. For exam ple, there is one study of genetic changes in sets of twins who developed T cell leukem i a at 9 years of age (reviewed in Sm ith and Zhang, 1998). Because of their sm all size, increased activity, and increased ventilation rates, as com p ared with those of adults, children m a y have greater exposure to benzene in the air on a unit-body-weight basis (U.S. EPA, 1998a). Infants and children also

109m a y be vulnerable to leukem i a induction from benzene because their hem a topoietic cell populations are undergoing m a turation and differentiation (U.S. EPA, 1998a).

4.4.2. Gender Differences

Hum a n exposure data regarding differential effects of benzene exposure on m a les and fem a les are very lim ited. Som e studies indicate that fem a les m a y be at greater risk than m a les. Sato et al. (1975) exposed m a le and fem a le workers to 25 ppm (80 m g /m 3 ) benzene for 2 hours and showed that benzene was retained longer in the fem a le subjects. During 5 hours of exposure, the blood concentrations were higher in m e n, whereas at the end, exhalation breath levels of both m e n and wom e n were essentially equivalent. However, 4 hours after the exposure had stopped, the fem a le subjects’ blood and exhaled air benzene levels were higher than those of

m a le subjects. The shape of the decay curve was significantly steeper in the m a les. Recently, Brown et al. (1998), using the Sato et al. (1975) data, developed a PBPK m odel sim u lation for adult m e n and wom e n. Results dem onstrated that physiochem ical gender differences result in

wom e n m e tabolizing 23–26% m o re benzene than do m e n when subjected to the sam e exposure conditions. Benzene blood levels were generally higher in m e n. These results suggest that

wom e n m a y be at significantly higher risk for certain effects of benzene exposure. In the occupational benzene exposure study by Rothm an et al. (1996a), 21 of the 44 workers in both the exposed and control groups were fem a le. The study, however, did not indicate that either gender

was m o re affected.

There are lim ited indications of gender differences in studies with test anim als. The m o st frequently observed gender difference is a greater sensitivity of m a le m i ce to benzene. W a rd et al. (1985) reported that m i croscopic exam ination revealed that changes in the thym us, bone

m a rrow, lym ph nodes, spleen, and reproductive organs occurred m o re often and with greater severity in m a les than in fem a les following inhalation exposure at 300 ppm (958 m g /m 3 ) of CD-1

m i ce for 13 weeks. Male DBA/2 m i ce showed greater depression in bone m a rrow cellularity than did fem a le DBA/2 m i ce in response to inhalation exposure to 300 ppm (958 m g /m 3 ) for 13 weeks (Luke et al., 1988a).

Male m i ce were also m o re sensitive to benzene by oral exposure (NTP, 1986). In a 17- week oral gavage study, the LOAEL for hem a tologic effects in m a le B6C3F1 m i ce was 50

m g /kg, com p ared with 400 m g /kg for fem a les. Fem a le rats, in contrast, appear to be m o re sensitive than m a les to benzene toxicity. In the NTP (1986) 17-week oral gavage study, the LOAEL for hem a tologic effects for fem a le F344/N rats was 25 m g /kg, com p ared with 100
mg / k g fo r ma l e s

.

Male m i ce are twofold to threefold m o re sensitive to the genotoxic effects of benzene, as

m easured by m i cronuclei induction and sister chrom atid exchanges (Luke et al., 1988a). Available studies indicate that the differential susceptibility to benzene is subject to horm onal regulation of the CYP2E1. Renal levels of CYP2E1 in m a les are 20-fold higher than in fem a les, but studies have indicated that renal levels of CYP2E1 can be induced in fem a les by testosterone
treatm e nt (Hu et al., 1993; Pan et al., 1992). Ke nyon et al. (1996) found that m a le B6C3F1 m i ce have an alm o st twofold faster rate of benzene oxidation than do fem a les. Phenol disappearance from the blood was also faster in m a le m i ce, suggesting that phenol m e tabolism is faster in

m a les. These differences in benzene m e tabolism correlated with the sensitivity to genotoxicity.

110Corti and Snyder (1996) found that CFU-E levels were depressed in livers of m a le fetuses of Swiss-W e bster m i ce exposed in utero to 10 ppm (32 m g /m 3 ) benzene in com p arison with controls. Sim ilarly, bone m a rrow CFU-E was also depressed in 6-week-old adult m a le

m i ce that had been exposed in utero. No depression in CFU-E was observed in fem a le fetuses or adult fem a les. This suggests that differences in the sensitivity of m a le m i ce 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 im m a ture m a le and fem a le m i ce did not differ in sensitivity to benzene.

Green et al. (1978) reported that exposure of pregnant m i ce to 2200 ppm (7028 m g /m 3 ) benzene caused a greater increase in m i ssing sternebrae in fem a le fetuses than in m a le fetuses, but this was not observed at concentrations of 300 ppm (958 m g /m 3 ) or less. The dose causing this effect induced extrem e m a ternal toxicity, com p licating interpretation of this effect. The apparently increased sensitivity of the fem a le fetus to the effects of benzene is supported by other observations that fem a le rabbits, m i ce, and rats are m o re 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 m a les and fem a les m a y be related to horm onal differences.

4.4.3. Genetically Susceptible Populations

Significant sources of variability in the population stem from genetic polym orphism s in key enzym e s involved in the m e tabolism of benzene, nam e ly, CYP2E1, NADPH-dependent quinone oxidoreductase, MPO, GSH transferase, and others. Dietary and endogenous sources of phenol, hydroquinone, and other prim a ry m e tabolites of benzene confer potentially large differences in susceptibility to benzene toxicity. These polym orphism s m a y increase or decrease an individual’s susceptibility to the toxic effects of benzene, as described below.

Sm art and Zannoni (1984, 1985) dem onstrated the im portance of reductase activity in bone m a rrow as a protective m echanism against benzene toxicity. Recent studies have also shown the im portance of polym orphic form s of GS H transferase as a protective characteristic. By the sam e token it appears that the reason that the liver is not a target for benzene toxicity is its high level of reductase activity, com p ared with the relatively low level in the bone m a rrow, where oxidative reactions predom inate. Furtherm ore, London et al. (1997) described a polym orphism in MPO, an enzym e im portant in activation of polyhydroxylated m e tabolites of benzene to quinones, suggesting that people who genetically display lower levels of this enzym e

m a y display a lower risk to carcinogenesis. Although biom arkers for susceptibility to benzene have not been validated, it m i ght be hypothesized that an individual who displays high CYP2E1 activity and low GSH transferases in liver and/or bone m a rrow and low bone m a rrow 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) m a y play a critical protective role in benzene toxicity. This enzym e detoxifies the reactive 1,4-benzoquinone generated by MPO by reducing it back to hydroquinone. Ross (1996)

sum m a rized 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 m u tation in the NQO1 gene that leads to a total loss of NQO1 activity. This appears to be a true polym orphism in the NQO1 gene, because Rosvold et al. (1995) found that the frequency of this m u tant allele in a

111reference population was 13% and Edwards et al. (1980) reported that NQO1 was absent in 4% of sam p les taken from a British population. If the postulated role of NQO1 in detoxification of benzene m e tabolites is correct, then individuals who lack this enzym e activity could be especially susceptible to benzene toxicity. Rothm an et al. (1997) carried out a case-control study in a population of benzene-exposed workers in Shanghai, China. CYP2E1 and NQO1 genotypes were determ ined by polym erase chain reaction-restriction fragm ent length polym orphism analysis, and CYP2E1 activity was estim ated by the fractional excretion of chlorzoxanone. Investigators found that subjects who had both rapid chlorzoxanone excretion and two copies of the NQO1 609 C 

###  T m u tation had a 7.6-fold increased risk of hem a totoxicity when com p ared 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 hem a totoxicity. Those individuals who had rapid CYP2E1 activity and heterozygous or hom ozygous variant NQO1 genotypes were at greatest risk of developing benzene-induced hem a totoxicty.

Seaton et al. (1994) reported differences of up to 13-fold in liver m i crosom e CYP2E1

am ong individual hum ans. Hepatic m e tabolism of benzene by the CYP2E1 enzym e has been shown to be the first step in generation of reactive benzene m e tabolites that are responsible for the toxicity of benzene. Thus, differences in CYP2E1 between individual hum ans could indicate potential differential susceptibility to benzene toxicity.

Rossi et al. (1999) investigated the role of genetic polym orphism s in m odulating urinary excretion of two benzene m e tabolites—MA and S-phenylm e rcapturic acid—in 59 nonsm oking city bus drivers, who were professionally exposed to benzene via vehicle exhausts. Genetic Table 9. Joint effects of CYP2E1 activity and NQO1 genotype on benz ene-induced hematotoxicity in Chinese Workers

CYP2E1 activity NQO1 genotype

Slow W ild type

Slow Variant
Rapid W ild type
Rapid V ariant Cases

8

6

21

13 Odds ratio (95% CI) benz ene hematotoxicity a

1.0
2.4 (0.6–9.7)
2.9 (1.0–8.2)
7.6 (1.8–31.2)

a Adjusted for the m a tching variables age and sex.

Source: Rothm an et al., 1997.

polym orphism s at six loci encoding cytochrom e P450-dependent m onooxygenases (CYP2E1 and CYP2D6), GSH transferases (GSTT1, GSTP1, and GSTM1), and NQOR were determ ined by

112polym erase chain reaction based m e thods. Metabolic variability was observed for CYP2D6, GSTT1, and NQOR, potentially contributing polym orphism as one of the risk factors for benzene-induced adverse health effects. However, no evidence em erged for a possible role of CYP2E1, GSTM1, and GSTP1 polym orphism s in determ ining the wide differences observed in the biotransform ation.

4.5. HAZ ARD IDENTIFICATION SUMMARY

Studies in experim e ntal anim als and hum an occupational studies have pointed to the changes in peripheral blood and bone-m arrow-induced exposure to benzene as being am ong the

m o st sensitive indicators of the com pound’s toxicity. Principal am ong the adverse effects described are leukopenia, lym phocytopenia, granulocytosis, anem ia, and reticulocytosis.

Hum a n occupational exposure studies have revealed clear evidence of benzene-induced hem a totoxicity (Table 1). Difficulties arise in directly applying m o st of these findings to quantitative estim ation of potential hum an health im pacts. This is due to uncertainty in defining the level of benzene exposure and to potential exposure to m i xtures of other potentially harm ful substances. One hum an occupational study, however, reported evidence that can be used to support the derivation of RfC and RfD values for benzene. In a sm all cross-sectional study of 44 age- and gender-m atched controls, Rothm an et al. (1996a) observed a dose-response relationship and inverse correlation between hem a tologic responses and exposure level. Six blood
param eters (ALC, W BC count, RBC count, HCT, platelets, and MCV) were significantly different in the high-benzene exposure group (8-hour TW A of 91.9 ppm [294 m g /m 3 ]) in

com p arison with controls. However, in the lower-exposure group (m edian 8-hour TW A of 13.6
ppm [43.4 m g /m 3 ], only ALC, RBCs, and platelets were significantly different.

In a second case-control study of a rubber worker cohort em ployed between 1939 and 1976 that relied on controversial exposure estim ates (W ard et al., 1996), a strong exposureresponse relationship between W BC count and estim ated 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 m a xim u m daily exposure estim ate was 34 ppm (109
mg / m 3 ). There was no evidence for a threshold, suggesting that exposure to relatively low levels of benzene (e.g., < 5 ppm [16 m g /m 3 ]) could result in hem a tologic suppression, if the exposure estim ates are correct.

Further support for the hypothesis that low exposures to benzene can reduce lym phocyte
counts com e s from a recent study (Bogadi-Sare et al., 2000) of fem a le em ployees in the shoem a king industry. This study found that the num ber of circulating B lym phocytes was lower in 49 shoe workers exposed to benzene concentrations lower than 15 ppm (48 m g /m 3 ) as

com p ared with 27 nonexposed controls. The authors concluded that benzene concentrations lower than 15 ppm (48 m g /m 3 ) can induce depression of circulating B lym phocytes. Additional evidence that low exposures to benzene induce reductions of blood param e ters com e s from a study of workers em ployed in factories in China (Dosem eci et al., 1996). Significant relative risks of “benzene poisoning” were recorded in workers exposed to as little as 5–19 ppm (16–61
mg / m 3 ) during the m o st recent 1.5 years as com p ared with workers exposed to < 5 ppm (16
mg / m 3 ). The authors defined benzene poisoning by two criteria: a W BC count of < 4000 cells/m m 3 and a platelet count of < 80,000/m m 3

. These data suggest that a threshold of

113hem a totoxicity in hum ans is observable in the 5–19 ppm (16–48 m g /m 3 ) range of exposure to benzene.

Hem a totoxicity has been observed in experim e ntal anim al studies following both oral and inhalation exposure (Table 6). Hem a tologic abnorm alities developed rapidly and were evident after only 2 weeks of exposure, but there was com p aratively little progression with longer exposure periods (W ard et al., 1985; Cronkite et al., 1985). The dose-response studies of W a rd et al. (1985) identified a NOAEL of 30 ppm (96 m g /m 3 ) and a LOAEL of 300 ppm (958 m g /m 3 ) in CD-1 m i ce. The results of Cronkite et al. (1985, 1989) suggest that peripheral blood elem ents

m a y have the capacity to recover from benzene-induced toxicity, depending on the duration of exposure and concentration. For exam ple, lym phocyte counts, which were depressed in C57BL/6 m i ce exposed to benzene at 300 ppm (958 m g /m 3 ) for 2–16 weeks, tended towards

norm al values during a 4- to 8-week recovery period.

Benzene-induced peripheral blood abnorm alities reflect a disruption at all levels of hem a topoiesis in the bone m a rrow (Toft et al., 1982; Snyder, 1987; Cronkite et al., 1985; Seidel et al., 1989; MacEachern et al., 1992). Bone m a rrow cellularity, stem cell com p artm ents, granulocytic and erythropoietic progenitor cells, and bone m a rrow m acrophages have all been observed to be adversely affected following benzene exposure. Recent evidence has suggested that the inhibitory effect of benzene on lym phocyte proliferation can be m e diated through the suppression of cytokine production (Renz and Kalf, 1991). The fact that both in vivo and in vitro studies have dem onstrated the ability of som e m e tabolites to induce toxicological effects qualitatively sim ilar to those of benzene provides evidence consistent with the concept that the hem a totoxicity of benzene m a y be m e diated through its m e tabolites.

There is no convincing evidence that benzene produces either reproductive or developm ental toxicity in hum ans. The available data are sum m a rized in Tables 4 and 5. Most studies consisted of sm all num bers of subjects, lacked im portant experim e ntal details, involved (in alm o st all cases) concom itant exposure to other chem icals, and did not provide m onitoring data or quantitative dose-response inform ation.

A num ber of reproductive and developm ental studies have been conducted in a wide range of experim e ntal anim als (Tables 7 and 8). Changes in testicular weight in guinea pigs (with slight histopathologic alterations) and rats were observed by W o lf et al. (1956), and reductions in ovarian and testicular weights with histopathologic alterations were observed in

m i ce by W a rd et al. (1985) after prolonged exposure to benzene. However, the concentrations were higher than those observed to cause hem a totoxicity. Several studies have suggested that benzene can cause developm ental toxicity in the absence of m a ternal 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 developm ental toxicity, the LOAEL and NOAEL values associated with reproductive/developm ental toxicity are higher than for hem a totoxicity.

The hem a tologic effects in offspring of dam s exposed to benzene during gestation were observed at exposure concentrations sim ilar to those in adult anim als (Keller and Snyder, 1986, 1988; Corti and Snyder, 1996). Changes in granulopoiesis and hem a topoiesis in offspring at

114som e growth stages were observed. Thus, the results were not consistent. The results of these studies cannot be used for quantitative evaluation of hum an health risk because of the lim ited

num ber of replicate anim als.

Sym p tom s of neurotoxicity have been observed in hum ans following acute exposure to relatively high concentrations of benzene (Brief et al., 1980; Kraut et al., 1988). Kahn and Muzyka (1973) reported subjective sym p tom s —headache, fatigue, difficulty sleeping, and

m e m o ry loss—in factory workers exposed to benzene. Baslo and Aksoy (1982) reported neurological abnorm alities in patients diagnosed with aplastic anem ia after prolonged exposure to benzene. These reports, however, have obvious deficiencies: lack of exposure data, sm all

num bers of subjects, and unknown exposure to other chem icals. Thus, no reliable quantitative evaluation of neurotoxicity in hum ans is possible.

Neurotoxicity studies with experim e ntal anim als were also very lim ited. 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 m onoam i ne neurotransm itters in the brain of rats at doses of 8 and 40 m g /kg/day, but further increases were not observed at 180 m g /kg/day. Benzene-induced behavioral and learning disorders in m i ce following inhalation exposures of 100 ppm (319 m g /m 3 ) or greater for 5 days (Evans et al., 1981; Dem p ster et al., 1984). Li et al. (1992) observed increased grip strength and enhanced rapidresponse m aze perform ance after exposure to 0.78 ppm (2.5 m g /m 3 ) for 30 days and decreases in these param e ters at higher concentrations. There were, however, several lim itations in the experim e ntal procedures used in these experim e nts. There is a lim ited body of evidence indicating that benzene is neurotoxic; however, there are no hum an or anim al studies that could be used for quantitative evaluation of potential hum an health risks.

No com p rehensive im m unotoxicity studies in hum an populations have been reported, but the study by Rothm an et al. (1996a) indicates decreased ALC as the m o st sensitive indicator of benzene toxicity. This could be interpreted as an effect on im m une function.

Im m unotoxicity studies in experim e ntal anim als have dem onstrated dose-related adverse effects on spleen weight and cellularity and on several m easures of im m une function following both oral and inhalation exposure to benzene (Hsieh et al., 1988b; Fan, 1992; W h ite et al., 1984; Rosenthal and Snyder, 1985; Rozen et al., 1984). NOAEL and LOAEL values for adverse effects on im m une function were sim ilar to those established for hem a totoxic effects. These studies indicate that even short-term exposures to benzene adversely affect the im m une response in experim e ntal anim als.

There is no convincing evidence to indicate that children are m o re susceptible to the toxic effects of benzene; however, there is evidence that differences in gender and subpopulation susceptibility m a y exist. Differences in responsiveness to benzene have been observed am ong species. Mice were found to be m o re sensitive than rats (W ard et al., 1985; Snyder et al., 1978, 1980). Intraspecies variability has also been dem onstrated. AKR/J m i ce were found to be m o re sensitive than C57BL/6 m i ce (Snyder et al., 1978, 1980). Absorption studies with hum ans suggest that absorption by fem a les is higher than by m a les, and m odeling results indicate that fem a les m e tabolize 23–26% m o re benzene than m e n under the sam e exposure conditions (Sato et al., 1975; Brown et al., 1998). Differences in benzene m e tabolism could result in differences

115in susceptibility, because conversion of benzene to m e tabolites is necessary for expression of toxicity. Several studies have indicated that m a le m i ce are m o re sensitive than fem a les (W ard et al., 1985; NTP, 1986; Siou et al., 1981; Kenyon et al., 1996; Corti and Snyder, 1996). In contrast, other studies using m i ce, rats, and rabbits have indicated that fem a les are m o re sensitive than m a les (Desoille et al., 1961; Ito, 1962a-d; Sato et al., 1975).

The balance of NQO1 NAD(P)H activity to peroxidase activity in cells m a y be im portant in determ ining benzene toxicity by m odulating the concentration of 1,4-benzoquinone, a suspected toxic m e tabolite. A m u tant allele for lack of NQO1 NAD(P)H activity has been reported in a hum an 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. Rothm an et al. (1997) also observed that Chinese workers hom ozygous for the m u tant allele and with high chlorzoxazone excretion (a m easure of CYP2E1 activity) had a 7.6-fold higher risk of developing benzene toxicity. Thus, there is good experim e ntal evidence to indicate that benzene-sensitive hum an subpopulations m a y exist.

Several experim e ntal anim al studies have observed that ethanol and benzene have an interactive effect on the production of hydroxylated benzene m e tabolites (Nakajim a et al., 1987). W h en adm i nistered at the sam e tim e, ethanol and benzene enhanced the num ber of erythroid progenitor cells in C57BL/6J m i ce (Baarson and Snyder, 1991), decreased the num ber of CFU-E cells per fem u r in BDF1 m i ce (Seidel et al., 1990), and produced a 70% reduction of CFU-E in

m a le Swiss-W e bster m i ce livers (Corti and Snyder, 1996). On the other hand, a 40% elevation of CFU-E occurred in fem a le Swiss-W e bster m i ce exposed to benzene and ethanol com p ared 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 hum an population.

5. DOSE-RESPONSE ASSESSMENTS

The prim a ry target organs for benzene toxicity are the hem a topoietic, im m une, and nervous system s. There are several reports of short-term exposure in hum ans and anim als. Data on the effects of benzene on the hum an hem a topoietic system following occupational inhalation exposure are scant, but they indicate effects such as leukopenia, anem ia, and throm bocytopenia. Data on hem a tologic effects in experim e ntal anim als 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., Aoyam a , 1986; Li et al., 1986; Green et al., 1981a), and in bone m a rrow cells (e.g., Cronkite et al., 1989; Dem p ster and Snyder, 1991). However, these studies were generally determ ined to be unsuitable for derivation of the RfD/RfC because they involved short-term exposures and sm all num bers of experim e ntal anim als; in addition, exposure levels were often too high to establish m eaningful LOAELs or NOAELs.

Data on adverse hem a tologic effects in hum ans following interm ediate 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). Howe ver, in the m a jority of these studies the exposure level estim ates and duration of exposure are poorly defined and/or characterized, and potential confounding is problem atic; therefore, these studies could not be used in determ inating

116the RfC and RfD, with the exception of the Rothm an et al. (1996a) study, which is used as the key study. The study of occupational inhalation exposure to benzene by Tsai et al. (1983) seem s to suggest a NOAEL for hem a tologic 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 em ployed survivor population. No relevant hum an data are available to evaluate hem a tologic effects following oral exposure.

Several subchronic and chronic studies in experim e ntal anim als following inhalation exposure (see Table 6) and oral exposure (NTP, 1986; Huff et al., 1989; W o lf et al., 1956; see also Section 4.4.2.1) identified NOAELs and LOAELs for hem a tologic effects and were considered in the derivation of the RfC and RfD. Ultim ately, the results of the W a rd et al. (1985) subchronic inhalation study and the NTP (1986) chronic gavage study were selected as the m o st appropriate anim al data for deriving an RfC and an RfD. These RfC and RfD values are calculated for com p arison with the values derived from the Rothm an et al. (1996a) hum an study.

W h ether the hem a totoxic/im m unotoxic effects of benzene and its carcinogenic effects are due to a com m on m echanism is not yet known. This is in part due to the fact that although the bone m a rrow depressive effects of benzene in hum ans can be readily duplicated in several experim e ntal anim al m odel system s, a suitable experim e ntal anim al system for the induction of leukem i a has not yet been developed. In addition, the hem a totoxicity/im m unotoxicity of benzene leads to significant health effects apart from potential induction of leukem i a. Although the decreased ALC and leukem i as observed from benzene exposure both result from bone

m a rrow toxicity, they do not necessarily result from the sam e m echanism s , and decreased ALC

m a y not be a necessary precursor for leukem i a. For exam ple, decreased ALC m a y be the result of cytotoxicity independent of the interaction of benzene m e tabolites with DNA and/or DNAassociated proteins and tum o r form ation. Thus, integration of the cancer and noncancer health

assessm ents of benzene is not considered possible given the current state of knowledge regarding

m ode of action.

5.1. INHALATION REFERENCE CONCENTRATION (RfC)

5.1.1. Choice of Principal Study and Critical Effect

The hum an occupational inhalation study by Rothm an et al. (1996a) was selected as the principal study for the derivation of the RfC because it is a well-conducted hum an exposure study that dem onstrates a dose-response relationship for hem a tologic effects, responses that are considered to be am ong the m o re 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 m g /L, because no increased hem a tological changes were observed am ong U.S. refinery workers chronically exposed to an average 0.53 ppm of benzene. This value was

assum ed 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 num bers were sim ilar to those of the U.S. EPA, although different studies were used. The cross-sectional Rothm an et al. (1996a) study provides exposure-response data for som e 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 m odeling to derive a point of

117departure for the calculation of the RfC. Reduced ALC is the m o st sensitive of the hem a tologic effects reported in this study, and it was selected as the critical effect. For com p arison purposes, a chronic inhalation RfC is also derived from the subchronic experim e ntal anim al study of W a rd et al. (1985), based on the critical effect of decreased HCT (see Section 5.1.5).

Rothm an et al. (1996a) com p ared the hem a tologic outcom e s of 44 workers occupationally exposed to benzene with those of 44 age- and gender-m atched unexposed controls, all based in Shanghai, China. ALC, W BCs, RBCs, and platelets were all significantly decreased and MCV was significantly increased in the exposed group (m edian 8-hour TW A of 31 ppm [99 m g /m 3 ]) as com p ared to those in the control group. These effects are consistent with the hem a totoxic effects of benzene shown in Aksoy (1989), Goldstein (1988), and Dosem e ci et al. (1997). In the low-dose subjects (< 31 ppm [99 m g /m 3 ], m e dian 8-hour TW A of 13.6 ppm [43.4 m g /m 3 ]), the ALC, RBCs, and platelet count were reduced com p ared with controls. Sim ilarly, in a selected subgroup exposed to a m e dian 8-hour TW A of 7.6 ppm (24 m g /m 3 ) benzene, a statistically significant difference in ALC versus controls was observed ( p <0.03).

The study by Rothm an et al. (1996a) is notable am ong epidem iology studies because benzene exposures were m onitored; exposure to other chem icals, including toluene, was

m i nim a l; and subjects were com p ared with m a tched controls. Furtherm ore, a dose-response relationship was established between ALC and benzene exposure, as m onitored by organic vapor passive dosim etry and the level of benzene m e tabolites in the urine. As discussed in Section 4.1.2.1, the m e dian 8-hour TW A of 7.6 ppm (24 m g /m 3 ) was designated the LOAEL for these effects and is used herein to calculate a chronic inhalation RfC for benzene hem a totoxicity in hum ans for com p arison with the RfC derived by BMD m odeling of the ALC exposure-response data.

The choice of a reduction in ALC as the prim a ry effect in an RfD/RfC derivation is partially based on the response’s potential role as a “sentinel” effect for a cascade of early hem a tologic and related biological changes that m i ght be expected to result in the m o re profound

exam ples of benzene poisoning observed in other cohorts of the National Cancer Institute/Chinese Academ y of Preventive Medicine study, as described by Dosem eci et al. (1996). Considered together, the statistically significant inverse correlation between ALC and the level of benzene m e tabolites in the urine of exposed subjects in the Rothm an et al. (1996a) study and the strong relationship between exposure to benzene in persons with benzene poisoning observed by Dosem eci et al. (1996) point clearly to the utility of fluctuations in ALC

as a m a rker for the onset of potentially harm ful hem a tologic changes. That ALC depletion is

accom p anied by gene-duplicating m u tations in som a tic cells under the sam e range of exposure conditions suggests that benzene can cause repeated dam a ge to longer-lived stem cells in hum an bone m a rrow, further im plicating the com pound as etiologically im portant in the onset of benzene-associated leukem i a. These findings underline the im portance of basing public health concern for benzene on a toxicological effect that is representative of the earliest biological changes induced by the com pound.

In sum m a ry, there is overwhelm ing evidence, in both experim e ntal anim als and hum ans, in the published literature cited in this report and in the cancer update docum ents (U.S. EPA, 1998a, 1999a), that chronic exposure to benzene leads to the onset of irreversible bone m a rrow

118depression, which is characterized by anem ia, leukocytopenia with em phasis on lym phocytopenia, and/or throm bocytopenia.

It is a well-known fact that benzene is a leukem ogen. The m echanistic steps that lead to that outcom e involve decreases in circulationg blood cells, as characterized by abnorm al m a rrow architecture, inadequate hem a topoiesis, and dem onstration of chrom osom al dam a ge in m a ny cells. Benzene toxicity is a continuum of events, leading from decreases in circulating blood cells to pancytopenia and aplastic anem ia or to MDS and acute nonlym phocytic leukem i a. Although the role of several m e tabolites of benzene and the “dose x tim es” relationship are only partially understood, the ultim ate result as a function of m a rrow dysplasia or aplasia or leukem i a is undisputable. In the Rothm an et al. (1996a) study, benzene air levels were inversely correlated with the absolute lym phocyte counts am ong exposed workers. This has been the subject of m o re detailed discussion in several sections of this docum ent as well as in the cancer update docum ents (U.S. EPA, 1998a, 1999a). The reader’s attention is drawn to sections 4.1, 4.2, 4.5, as well as to som e selected references: Rothm an et al. (1996a, b), Dosem eci et al. (1997), Bogardi-Sari et al. (2000), Hsieh et al. (1988b), NTP (1986), and W a rd et al. (1985).

5.1.2. Benchmark Dose Modeling

The exposure-response data from Rothm an et al. (1996a, Table IV) that were used for the BMD m odeling are reproduced below.

Median exposure ALC Transform ed (ppm ; 8-hr TW A) Num b er of subjects (m ean ± SD × 10 3 /μL blood) exposures 0.02 (control) 4 4 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 m odeling was done using EPA’s Benchm ark 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 m odels, the exposure levels were first transform ed according to the equation d 

### 1 = ln(d + 1). Then the exposure-response data were fit using the continuous linear m odel, which has the form

Y[dose] = 

###  0 + 

###  1 *dose,

where, in this case, Y[dose] is the m ean ALC and dose is the transform ed exposure, d 

### 1

. The
param e ters were estim ated using the m e thod of m a xim u m likelihood. A constant variance m odel was used. The resulting param e ter estim ates are:

Variable Estim ate SE

119

###  0 1.91029 26.8966 

###  1 -0.129822 70.7846

The m odel 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 benchm ark response (BMR) of one standard deviation change from the control m ean was selected, as suggested in EPA’s draft Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b). This default definition of a benchm ark response for continuous endpoints corresponds to an excess risk of approxim a tely 10% for the proportion of individuals below the 2 nd percentile (or above the 98 th percentile) of the control distribution for norm ally distributed effects (see U.S. EPA, 2000b). The software uses the estim ated standard deviation. For the resulting BMC, a 95% lower confidence lim it (BMCL) was calculated using the likelihood profile m e thod. A BMC of 2.69 and a BMCL of 2.10 were obtained with the transform ed
exposures. Transform ing these values back to the original exposure scale yields the following values: BMC = 13.7 ppm , 8-hour TW A; BMCL = 7.2 ppm , 8-hour TW A.

A two-degree restricted polynom ial m odel also adequately fits the data, yielding a BMC of 3.00 and a BMCL of 2.14 with the transform ed doses (i.e., 19.1 and 7.5 ppm 8-hour TW A, respectively); however, the linear m odel was selected because it is the m o st parsim onious (i.e., it

120Figure 2. Linear model of ALC data from Rothman et al. (1996a). Figure 2. Linear model of ALC data from Rothman et al. (1996a).

is the m odel with the fewest param e ters that fits the data) and 2 × (the absolute difference in the log-likelihood values between the two m odels) < 

### 3 1 2 , that is, the param eter 

###  2 for the coefficient of dose 2 in the two-degree polynom ial m odel is not statistically different from 0. Results using a restricted power m odel were identical to those using the two-degree polynom ial.

Note that the 7.6 ppm exposure subgroup (11 subjects) was not included explicitly in the
exposure-response m odeling because it was a subset of the 13.6 ppm group (22 subjects) and the data were not available for the rem a ining 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 m eant that som e of the data, for one-fourth of the exposed subjects, would not have been used at all. Com parison analyses were conducted to exam ine the effects of using the 7.6 ppm group and either om itting the subjects in the 2 nd exposure quartile or estim ating the exposure level of the 2 nd quartile as the m i dpoint between 13.6 and 31 ppm

. The BMCLs were sim ilar 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 docum ent (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 m g /m 3 and adjusting the 8-hour TW A occupational exposure to Table 10. Results of BMC modeling of Rothman et al. (1996a) data on benz ene and ALC a

121Dataset BMC (ppm) BMCL (ppm)
without 7.6 ppm subgroup 13.7 7.2
with 7.6 ppm subgroup,

om itting 2 nd quartile 14.7 7.4

with 7.6 ppm subgroup, estim ating 2 nd quartile 17.1 8.4

a Using log-transform ed exposures, a linear m odel, and a BMR level of one standard deviation change from the control m ean.

an equivalent continuous environm ental exposure. The BMCL is first converted to m g /m 3 using the m o lecular weight of 78.11 for benzene and assum i ng 25°C and 760 m m Hg:

BMCL (m g/m 3 ) = 7.2 ppm × 78.11/24.45 = 23.0 m g /m 3

The converted value is then adjusted from the 8-hour TW A to an exposure concentration adjusted for continuous exposure using the default occupational m i nute volum e (U.S. EPA, 1994).

BMCL ADJ = BMCL (m g/m 3 ) × (VEho/VEh) × 5 days/7 days

BMCL ADJ = 23.0 m g /m 3 × (10 m 3 /20 m 3 ) × 5 days/7 days = 8.2 m g /m 3

where:

BMCL ADJ = the BMCL dosim etrically adjusted to account for continuous exposure

BMCL = occupational exposure level (8-hour TW A)

VEho = hum an occupational default m i nute volum e (10 m 3 /8 hours)

VEh = hum an am bient default m i nute volum e (20 m 3 /24 hours).

To calculate an RfC using the BMCL ADJ value of 8.2 m g /m 3 , 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 judgem e nt 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

m easured in the blood, but it is not a frank effect, and there is no evidence that it is related to any functional im pairm e nt at levels of decrem ent near the BMR. For a m o re serious effect, a larger factor, such as 10, m i ght be selected.

122Second, a factor of 10 was used for intraspecies differences in response (hum an variability) as a m eans of protecting potentially sensitive hum an subpopulations. Third, a subchronic-to-chronic extrapolation factor was applied because the m ean 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 assum e d hum an life span of 70 years) which has been used by the Superfund program as a cut-off for deriving a subchronic hum an reference dose (U.S. EPA, 1989). However, because the m ean 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/developm ental 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 = BMCL ADJ ÷ UF = 8.2 m g /m 3 ÷ 300 = 3 × 10 -2 mg/m 3

5.1.4. Comparison Analysis Based on the LOAEL

A m e dian 8-hour TW A concentration of 7.6 ppm was designated the LOAEL for the Rothm an et al. (1996a) study for a com p arison calculation of a chronic inhalation RfC for the critical effect of reduced ALC. A LOAEL ADJ can be calculated, as described for the BMCL ADJ in Section 5.1.3, by converting to m g /m 3 and adjusting from the occupational ventilation rate and interm ittent work-week schedule to a continuous 24-hour exposure, 7 days/week:

LOAEL (m g/m 3 ) = 7.6 ppm × 78.11/24.45 = 24.3 m g /m 3

LOAEL ADJ = 24.3 m g /m 3 × (10 m 3 /20 m 3 ) × 5 days/7 days = 8.7 m g /m 3

To calculate an RfC using the LOAEL ADJ value of 8.7 m g /m 3 , 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 (hum an 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 = LOAEL ADJ ÷ UF = 8.7 m g /m 3 ÷ 1000 = 9 × 10 -3 mg / m 3

This value is in good agreem ent with the RfC of 3 × 10 -2 mg / m 3 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 experim e ntal anim al study of W a rd et al. (1985) for com p arison with the RfC of 3 × 10 -2 mg / m 3 based on the Rothm an et al. (1996a) hum an study. The W a rd et al. (1985) study was selected because it used a relatively long inhalation exposure duration and an adequate num ber of anim als, and it

123provided dose-response data. The study identified both a LOAEL of 300 ppm and a NOAEL of 30 ppm

. The investigators exposed m a le and fem a le CD-1 m i ce and Sprague-Dawley rats to 0, 1, 10, 30, or 300 ppm (0, 3.2, 32, 96, or 960 m g /m 3 ) benzene 6 hours/day, 5 days/week for 91 days and m easured various hem a tologic endpoints. The m a le m ouse appears to be the m o st sensitive sex/species in this study. The exposure-response relationships for the different hem a tologic endpoints were m odeled using a BMD m odeling approach, and decreased HCT was chosen as the critical effect. The exposure-response data from W a rd et al. (1985, Table II, 91 days) for the m a le m ouse are reproduced below.

Exposure Num b er of % HCT (ppm ) subjects (m ean ± 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 Benchm ark Dose Modeling Software (version 1.20) was used for the m odeling. An assum p tion of constant variance was used, although the test for hom ogeneity of the variances failed. The continuous linear, polynom ial, and power m odels all resulted in the sam e BMC and BMCL estim ates; however, the linear m odel had better results for the fit statistics. The linear

m odel had a p -value of 0.09, which is of borderline adequacy (EPA’s draft Benchmark Dose Technical Guidance Document [U.S. EPA, 2000b] recom m e nds a p -value of 

###  0.1), whereas the other m odels had p -values of 0.04. Thus the continuous linear m odel was selected. This m odel has the form

Y[dose] = 

###  0 + 

###  1 * dose

where, in this case, Y[dose] is the m ean HCT level and dose is the experim e nt exposure concentration. The param e ters were estim ated using the m e thod of m a xim u m likelihood. The resulting param e ter estim ates are

Variable Estim ate SE



###  0 40.0962 2.44042 

###  1 -0.0407075 329.232

See Figure 3 for a graphical display of the data and m odel.

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 m ean was selected, as suggested in the EPA’s draft BMD technical guidance docum ent (U.S. EPA, 2000b). The software uses the estim ated standard deviation. A 95% lower confidence lim it (BMCL) on the resulting BMC was calculated using the likelihood profile m e thod. A BMC of 100.7 ppm and a BMCL of 85.0 ppm

124Figure 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

125ppm ). RBC results were in between those for MCV and m ean corpuscular Hgb (MCH) and those for HCT and total Hgb; however, the m odel fits were not adequate for the RBC data; thus, the RBC results have m o re uncertainty.

To derive the RfC, the BMCL is used as the point of departure, as suggested in the draft BMD technical guidance docum ent (U.S. EPA, 2000b). An adjusted BMCL is calculated by converting ppm to m g /m 3 and adjusting from anim al experim e nt exposure to equivalent continuous exposure. (For conversion of the inhalation exposures across species, ppm equivalence was assum e d; this is identical to using EPA’s inhalation dosim etry m e thodology with the regional gas dose ratio for the respiratory tract region (RGDR
r ) = 1 [U.S. EPA, 1994].) The BMCL is first converted to m g /m 3 using the m o lecular weight of 78.11 for benzene and

assum i ng 25°C and 760 m m Hg:

BMCL (m g/m 3 ) = 85.0 ppm × 78.11/24.45 = 272 m g /m 3

The converted value is then adjusted to an equivalent continuous exposure as follows:

BMCL ADJ = 272 m g /m 3 × 6 hours/24 hours × 5 days/7 days = 48.5 m g /m 3

To derive the RfC, several UFs are applied to the BMCL ADJ . As discussed in Section5.1.3, 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/developm ental 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 hem a tologic fluctuations such as reductions in RBCs and W BCs in the high-dose m i ce were noted at interim sacrifice (14 days) as well as at
term ination (91 days), suggesting that the responses occurred early in the exposure cycle and then rem a ined com p aratively unchanged. Therefore, an overall UF of 3 × 3 × 10 × 3 × 3 = 1000 is used in the calculation of the chronic inhalation RfC from the BMCL ADJ , as follows:

RfC = BMCL ADJ ÷ UF = 48.5 m g /m 3 ÷ 1000 = 5 × 10 -2 mg / m 3

This value is in good agreem ent with the RfC of 3 × 10 -2 mg / m 3 calculated from the BMC from the Rothm an et al. (1996a) hum an study.

Sim ilarly, for com p arison purposes, a chronic inhalation RfC can be derived from the NOAEL of 30 ppm observed for hem a tologic effects in the W a rd et al. (1985) experim e ntal anim al study. First, the NOAEL is converted to m g /m 3 and adjusted to equivalent continuous exposure, as above:

30 ppm × 78.11/24.45 = 95.8 m g /m 3 95.8 m g /m 3 × (6 hours/24 hours) × 5 days/7 days = 17.1 m g /m 3

126(ppm equivalence across species was assum e d; this is identical to using EPA’s inhalation dosim etry m e thodology with RGDR
r =1 [U.S. EPA, 1994]). UFs are identical to those em ployed above, except that no NOAEL-to-LOAEL UF is used, thus the overall UF is 300, resulting in an RfC of

NOAEL ADJ ÷ UF = 17.1 ÷ 300 = 6 × 10 -2 mg / m 3

This value is also in good agreem ent with the RfC of 3 × 10 -2 mg / m 3 derived from the Rothm an et al. (1996a) study.

It should be noted, however, that other experim e ntal anim al studies have reported significant hem a tologic effects at benzene exposures of 10–25 ppm , which are lower than the NOAEL of 30 ppm from the W a rd et al. (1985) study. These studies have insufficient data for dose-response m odeling, and they used shorter exposure durations and/or fewer experim e ntal anim als than did the W a rd et al. study; nonetheless, they observed statistically significant hem a tologic effects at 10–25 ppm

. Baarson et al. (1984), for exam ple, exposed m a le C57BL/6J

m i ce (five/group) to 10 ppm benzene 6 hours/day, 5 days/week for 178 days and observed statistically significant reductions in blood lym phocytes at each of the three m onitoring tim e points (32, 66, and 178 days) as com p ared with controls. The m a gnitude of the reduction in lym phocytes ranged from about 53% at 32 days to about 68% at 178 days. Cronkite et al. (1985) exposed m a le and fem a le C57BL/6 BNL m i ce to various concentrations of benzene 6 hours/day, 5 days/week for 2 weeks and observed no decrease in blood lym phocytes at 10 ppm ; however, the investigators did observe a statistically significant reduction of about 21% at 25 ppm as

com p ared with controls (5–10 m i ce/group). Thus, lower RfCs than those calculated above for the W a rd et al. (1985) study are possible, based on other experim e ntal anim al results. In the

m ost extrem e case, using a LOAEL of 10 ppm and an overall UF of 3000 yields a LOAEL ADJ of 5.7 m g /m 3 and an RfC of 2 × 10 -3 mg / m 3

.

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 hum an occupational inhalation study of Rothm an 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 hum an exposure study that dem onstrated a dose-response relationship for hem a tologic effects, responses that are considered to be am ong the m o re sensitive indices of benzene toxicity (see Section 5.1.1 for m o re details). Furtherm ore, no relevant hum an data are available to evaluate hem a tologic effects following oral exposure. For com p arison purposes, a chronic oral RfD is also derived from the chronic experim e ntal anim al gavage study of the NTP (1986) on the basis of the sam e critical effect of reductions in lym phocyte 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 m odeling of the ALC data of Rothm an et al. (1996a) yielded a BMC of 13.7 ppm (8-hour TW A) and a BMCL of 7.2 ppm (8-hour TW A) for the

127default BMR of one standard deviation change from the control m ean response. This default definition of a BMR for continuous endpoints corresponds to an excess risk of approxim a tely 10% for the proportion of individuals below the 2 nd percentile (or above the 98 th percentile) of the control distribution for norm ally 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 TW A) is shown below.

The BMCL is first converted to m g /m 3 using the m o lecular weight of 78.11 for benzene and assum ing 25°C and 760 m m Hg:

BMCL (m g/m 3 ) = 7.2 ppm × 78.11/24.45 = 23.0 m g /m 3

The converted value is then adjusted from the 8-hour TW A to an exposure concentration adjusted for continuous exposure using the default occupational m i nute volum e (U.S. EPA, 1994).

BMCL ADJ = BMCL (m g/m 3 ) × (VEho/VEh) × 5 days/7 days

BMCL ADJ = 23.0 m g /m 3 × (10 m 3 /20 m 3 ) × 5 days/7 days = 8.2 m g /m 3

where:

BMCL ADJ = the BMCL dosim etrically adjusted to account for continuous exposure

BMCL = occupational exposure level (8-hour TW A)

VEho = hum an occupational default m i nute volum e (10 m 3 /8 hours)

VEh = hum an am bient default m i nute volum e (20 m 3 /24 hours)

In the support docum ent for the benzene cancer assessm ent on IRIS (U.S. EPA, 1999a), EPA provided a sim p le m e thod for extrapolating benzene-induced cancer risk from the inhalation to oral route. The sam e m e thod is applied here for noncancer (hem a topoietic) effects. The m e thod is based on the relative efficiency of benzene absorption across routes of exposure, especially pulm onary 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 hum an inhalation absorption studies (Nom iyam a and Nom i yam a , 1974; Pekari et al., 1992; Srbova et al., 1950) and the study of Sabourin et al. (1987), which com p ared inhalation and oral absorption in rats and m i ce.

Sabourin et al. (1987) found that the retention of 14 C-benzene by rats and m i ce 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 m i ce as exposure concentration increased from 32 to 3,200 m g /m 3 (10 to 1,000 ppm ). In the sam e study, gastrointestinal absorption of benzene adm i nistered by gavage was > 97% for doses between 0.5 and 150 m g /kg body weight. At oral doses below 15 m g /kg, > 90% of the 14 C excreted was in the urine as

128nonethyl acetate-extractable m aterial. At higher doses, an increasing percentage of the orally adm i nistered benzene was exhaled unm etabolized. Thus, in the dose range represented by the BMCL from the study by Rothm an et al. (a, b), absorption of a com p arable oral dose was

assum e d to be 100%. See also U.S. EPA (1999a) for m o re details about the route-to-route extrapolation of benzene inhalation results to oral exposures.

To calculate an equivalent oral dose rate, the BMCL ADJ is m u ltiplied by the default inhalation rate, m u ltiplied by 0.5 to correct for the higher oral absorption, and divided by the standard default hum an body weight of 70 kg (U.S. EPA, 1988):

8.2 m g /m 3 × 20 m 3 /day × 0.5 ÷ 70 kg = 1.2 m g /kg/day

5.2.3. RfD Derivation

To calculate an RfD using the BMCL ADJ -equivalent oral dose rate value of 1.2

m g /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 m o re serious effect, a larger factor, such as 10, m i ght be selected. Second, a factor of 10 was used for intraspecies differences in response (hum an variability) as a m eans of protecting potentially sensitive hum an 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/developm ental 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 m g /kg/day ÷ 300 = 4 × 10 -3 mg/kg/day

.

Use of a m odifying factor of 3 to account for uncertainty in the route-to-route extrapolation was considered; however, it was deem ed unnecessary. The RfD is based on hum an data for a sensitive endpoint; thus, it was felt that the com posite UF of 300 provides sufficient protection.

1295.2.4. Comparison Analysis Based on the LOAEL

A m e dian 8-hour TW A concentration of 7.6 ppm was identified from the Rothm an et al. (1996a) study as a LOAEL for the critical effect of reduced lym phocyte 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 LOAEL ADJ is calculated by converting to m g /m 3 and adjusting from the occupational ventilation rate and interm ittent work-week schedule to a continuous 24- hour exposure, 7 days/week:

LOAEL (m g/m 3 ) = 7.6 ppm × 78.11/24.45 = 24.3 m g /m 3

LOAEL ADJ = 24.3 m g /m 3 × (10 m 3 /20 m 3 ) × 5 days/7 days = 8.7 m g /m 3

Next, an inhalation absorption rate of 50% was used to calculate the absorbed benzene dose, and oral absorption was assum e d to be 100%, as discussed in Section 5.2.2. To calculate an equivalent oral dose rate, the LOAEL ADJ is m u ltiplied by the default inhalation rate,

m

u ltiplied by 0.5 to correct for the higher oral absorption, and divided by the standard default hum an body weight of 70 kg (U.S. EPA, 1988):

8.7 m g /m 3 × 20 m 3 /day × 0.5 ÷ 70 kg = 1.2 m g /kg/day

To calculate an RfD using the LOAEL ADJ -equivalent oral dose rate value of 1.2

m g /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 (hum an 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 m g /kg/day ÷ 1000 = 1 × 10 -3 m g /kg/day.

This value is in good agreem ent with the RfD of 4 × 10 -3 m g /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 Rothm an et al. (1996a) study because it was decided that it was preferable to use the hum an data rather than the experim e ntal anim al data. The NTP (1986) study rather than the Hsieh et al. (1988b) study was used for the com p arison analysis because it was m u ch larger (50 vs. 5 anim als/dose group), it was conducted for 2 years (vs. 28 days) with blood m easurem e nts m a de at m u ltiple tim e points, and it exam ined both m a le and fem a le rats and m i ce (vs. just m a le m i ce). The Hsieh et al. (1988b) study provided a lower LOAEL (8 m g /kg/day vs. 18 m g /kg/day adjusted to 7 days/week); however, the lower LOAEL is not crucial, because the m a in analysis currently uses a BMD approach. Both studies yielded sim ilar BMDLs for decreased ALC (0.7 m g /kg/day for the m a le rat NTP data vs. 1.4 m g /kg/day for the Hsieh et al. data; both based on a linear m odel with transform ed doses and a BMR of one standard deviation change from the control m ean).

130A chronic oral RfD was also derived from the NTP (1986) experim e ntal anim al study for

com p arison with the RfD of 1 × 10 -3 m g /kg/day based on the Rothm an et al. (1996a) hum an study. In the NTP (1986) study, F344 rats and B6C3F1 m i ce of both sexes were adm i nistered benzene by gavage 5 days/week for 103 weeks (see also Section 4.2.1.2). Male rats (50/group) were adm i nistered doses of 0, 50, 100, or 200 m g /kg and fem a les (50/group) were adm i nistered doses of 0, 25, 50, or 100 m g /kg. B6C3F1 m i ce (50/sex/group) were adm i nistered doses of 0, 25, 50, or 100 m g /kg. Blood was drawn from 10 random ly preselected anim als per species/sex/dose group at 12, 15, 18, and 21 m onths and from all anim als at the term inal kill at 24 m onths. Additional groups of 10 anim als of each sex and species were adm i nistered benzene for 51 weeks at the sam e doses of the 2-year study, and blood was drawn at 0, 3, 6, 9, and 12

m onths. This study identified a LOAEL of 25 m g /kg for leukopenia and lym phocytopenia in fem a le F344 rats and m a le and fem a le B6C3F1 m i ce and of 50 m g /kg in m a le F344 rats. These were the lowest doses tested and, thus, no NOAEL was identified.

Reduction in lym phocyte count was selected as the critical effect, and attem p ts were

m a de to m odel the dose-response relationships using a BMD m odeling approach. However, the dose-response analysis of the NTP data was problem atic for a num ber of reasons. First, only the grouped data, not the individual anim al data, were provided. Second, as discussed in Appendix N of the NTP (1986) report, not all of the groups were com p arable for a variety of reasons, including experim e nt design, differences in bleeding m e thods and tim e of bleeding, and differences in m easurem e nt instrum entation. Third, the grouped data exhibited substantial variability, and baseline m easurem e nts (vehicle control groups) were not constant over tim e. Furtherm ore, for the rats, the strong dose effect in lym phocyte count was accom p anied by significant tem poral variability, which appears to be largely due to tem poral variability in the vehicle control group m eans. For these reasons, it was not possible to do a pooled analysis across tim e points. Instead, m odeling was perform ed for each dataset in two data groupings within which the datasets are com p arable (m onths 6 and 9; and m onths 12, 15, 18, and 21), and ranges of results are presented. Each of these datasets had at m o st 10 anim als/dose, so the doseresponse results are not very robust. The m a les of both species exhibited m o re dram a tic and consistent reductions in lym phocyte count than did the fem a les, but it was not clear a priori which species was m o re sensitive; therefore, dose-response analyses were perform ed for both the

m a le m ouse (NTP Table N8) and the m a le rat (NTP Table N4).

The continuous linear, polynom ial, and power m odels in EPA’s Benchm ark Dose Modeling Software (version 1.20) were used for the m odeling. The software estim ates the
param e ters using the m e thod of m a xim u m likelihood. Most of the data were supralinear (i.e., the

m a gnitude of the reductions in lym phocyte count decreased with increasing unit dose) and not

am enable to m odeling with restricted m odels (in this case, with param e ters constrained to be nonpositive), as suggested in EPA’s draft Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b). Therefore, two different strategies were attem p ted. The first was to use unrestricted polynom ial m odels and exclude the high-dose group when necessary to fit the

m odels, and the second was to transform the dose data before m odeling. Dropping the high-dose groups alone did not resolve the problem because there was still too m u ch supralinear curvature in the two lower dose groups to fit restricted m odels. Results are sum m a rized below in Tables 11 and 12, for untransform ed and transform ed data, respectively.

131For each dataset, the selected m odel was chosen on the basis of the lowest Akaike’s Inform ation Criterion value with consideration of the graphical display, as suggested in the draft BMD technical guidance docum ent. For selecting between m odels within a fam ily of m odels (e.g., between a linear and a two-degree polynom ial m odel), consideration was given to the loglikelihood values to evaluate the statistical significance of adding an extra param e ter. In som e cases there was no m odel that adequately fit the data. The draft technical guidance docum ent suggests a p -value of 0.1 for m odel fit. In addition, there was substantial variability in these data, but the variability appeared to be random and not am enable to m odeling. Therefore, constant variance was assum e d for all the m odels, although in som e cases the variances failed the test for hom ogeneity.

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 m ean response was selected as suggested in the draft BMD technical guidance docum ent. 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 estim ated “constant” standard deviation as the standard deviation for all the group m eans. The 95% lower confidence lim it (BMDL) on the resulting BMD was then calculated using the likelihood profile

m e thod.

Because m a ny of the datasets could not be m odeled using the available m odels without dropping the param e ter constraints and som e tim es even excluding the highest dose group (Table 11), a second m odeling approach was tried in which the dose levels were transform ed before the dose-response m odeling was conducted, as was done for the hum an data m odeled in Section 5.1.2. The dose levels were transform ed according to the form ula: transform ed dose = ln(dose + 1). Results of the dose-response m odeling using the transform ed doses are presented in Table 12.

The results in Table 12 suggest that the m a le rat is m o re sensitive than the m a le m ouse to lym phocyte count reductions from exposure to benzene in this NTP (1986) gavage bioassay, because the ranges of BMDs/BMDLs are substantially lower for the m a le rat, especially for year 2. The ranges for the m a le rat are also fairly tight, and the m odels selected provide good fits to all the m a le rat datasets. However, all but one of the calculated BMDs for the m a le rat are over an order of m a gnitude below the lowest exposure dose of 50 m g /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 im pacts of m odel selection and the uncertainties inherent in extrapolating to lower doses.

Nevertheless, data from two drinking water studies (Hsieh et al., 1988b; W h ite et al., 1984) provide support for selecting a BMD in this range. These two studies were of shorter duration and used fewer experim e ntal anim als than did the NTP (1986) study; however, they do provide dose-response data for BMD m odeling and they also have the advantage of being drinking water studies; thus the benzene exposure scenario is m o re relevant to hum an oral benzene exposures. In one study, Hsieh et al. (1988b) exposed m a le CD-1 m i ce (five/group) to 0, 8, 40, or 180 m g /kg/day benzene in drinking water for 28 days. Hem a tologic effects were

132Table 11. BMD modeling results of the NTP (1986) male mouse and male rat lymphocyte counts, w i th untransformed data

Dataset Model Variance homogeneity F it BMD a (mg/k g ) BMDL a (mg/k g )

Male mouse

6-m onth unrestricted 2-degree polynom ial OK yes p =0.13 20.52 12.96

9-m onth unrestricted 2-degree polynom ial no yes p =0.72 23.06 13.70

year 1 range 20.52–23.06 12.96–13.70

12-m onth unrestricted 2-degree polynom ial OK no b

p =0.005 16.68 b 11.41 b

15-m onth linear no yes, p =0.98 51.94 36.80

18-m onth linear no yes, p =0.16 35.50 26.92

21-m onth linear no yes, p =0.25 42.91 30.44

year 2 range 35.50–51.94 26.92–36.80

Male rat

6-m onth unrestricted 2-degree polynom ial OK yes p =0.56 15.70 11.79

9-m onth unrestricted 2-degree polynom ial no yes p =0.37 16.88 12.72

year 1 range 15.70–16.88 11.79–1272

12-m onth; w/o high-dose
group unrestricted 2-degree polynom ial OK yes c 10.51 7.3

15-m onth; w/o high-dose
group unrestricted 2-degree polynom ial OK yes c 7.56 5.56

18-m onth; w/o high-dose
group unrestricted 2-degree polynom ial no yes c 13.84 9.07

21-m onth; w/o high-dose
group unrestricted 2-degree polynom ial OK yes c 7.21 5.03

year 2 range 7.21–13.84 5.03–9.07

a Unadjusted anim al doses in m g /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 3 2 -residuals; however, there were insufficient degrees of freedom for calculation of a p -value.

133Table 12. BMD modeling results of the NTP (1986) male mouse and male rat lymphocyte counts, w i th transformed dose data

Dataset Model Variance homogeneity F it BMD a (mg/k g ) BMDL a (mg/k g )

Male Mouse

6-m onth 2-degree polynom ial O K borderline p =0.047 19.68 6.57

9-m onth linear no yes, p =0.35 9.07 4.05

year 1 range 9.07–19.68 4.05–6.57

12-m onth linear OK yes, p =0.30 3.74 2.32

15-m onth power no yes, p =0.31 47.46 18.55

18-m onth power n o borderline p =0.09 28.93 13.99

21-m onth power no yes, p =0.15 23.34 5.80

year 2 range 3.74–47.46 2.32–18.55

Male rat

6-m onth power OK yes, p =0.30 9.92 4.52

9-m onth linear no yes, p =0.11 3.71 2.30

year 1 range 3.71–9.92 2.30–4.52

12-m onth linear no yes, p =0.22 1.34 0.95

15-m onth linear OK yes, p =0.93 1.34 0.95

18-m onth linear no yes, p =0.22 2.73 1.74

21-m onth linear OK yes, p =0.54 1.69 1.10

year 2 range 1.34–2.73 0.95–1.74

a Unadjusted anim al dose in m g /kg, after transform ing the results back according to the form ula: dose = exp(transform ed dose)

 1.

observed at all exposure levels. BMD m odeling of the ALC yielded a BMD of 2.2 m g/kg/day and a BMDL of 1.4 m g /kg/day, based on a linear m odel with transform ed doses and a BMR of one standard deviation change from the control m ean, as above. In the second study, W h ite et al. (1984) exposed fem a le B6C3F1 m i ce to 0, 12, 195, or 350 m g /kg/day benzene in drinking water for 30 days. BMD m odeling of the ALC (five to six m i ce/group) resulted in a BMD of 11.6

m g /kg/day and a BMDL of 5.3 m g /kg/day (also based on a linear m odel with transform ed doses and a BMR of one standard deviation change from the control m ean, as above).

The results in Table 12 from BMD m odeling of the m a le rat ALC data from the NTP (1986) study show the lowest BMDL of about 1 m g /kg at three tim e points in the second year; therefore, this was selected as the point of departure for an RfD calculation. Adjusting for

134exposure 7 days/week yields a BMDL ADJ of 0.7 m g /kg/day. As discussed in previoussubsections, 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 = BMDL ADJ ÷ UF = 0.7 m g /kg/day ÷ 1000 = 7 × 10 -4 m g /kg/day

This RfD value is in reasonably good agreem ent (within an order of m a gnitude) with the RfD value of 4 × 10 -3 m g /kg/day derived from the Rothm an et al. (1996a) hum an inhalation study.

For com p arison purposes, a chronic oral RfD can be derived from the LOAEL of 25

m g /kg identified for hem a tologic effects in the NTP (1986) study (there was no NOAEL). Adjusting from 5-day to 7-day exposure yields a LOAEL ADJ of 18 m g /kg/day, which can be used to calculate an RfD for benzene as follows:

RfD = LOAEL ADJ ÷ UF = 18 m g /kg/day ÷ 3000 = 6 × 10 -3 m g /kg/day,

where the com b ined UF of 3000 is m a de up of com ponent 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 agreem ent with the RfD of 4 × 10 -3 m g /kg/day calculated from the BMD analysis of the Rothm an et al. (1996a) hum an data.

5.3. DOSE-RESPONSE SUMMARY

For the derivation of a chronic inhalation RfC, the hum an occupational inhalation study of Rothm an et al. (1996a) was selected as the principal study, and the inhalation study of W a rd et al. (1985) was selected as a supporting experim e ntal anim al study. These two studies provided the best dose-response data for quantitatively evaluating the potential hum an health risks due to inhalation of benzene. The cross-sectional study of Rothm an et al. (1996a) was designated as the principal study because it is a hum an exposure study with well-quantified benzene exposure estim ates and data on sensitive hem a tological endpoints. The study com p ared the hem a tologic evaluations of 44 workers occupationally exposed to benzene with those of 44 age- and gender-m atched unexposed controls, all based in Shanghai, China. ALC, W BCs, RBCs, and platelets were all significantly decreased, whereas MCV was significantly increased in the group of exposed workers (m edian 8-hour TW A of 31 ppm [99 m g /m 3 ]), com p ared with the age- and sex-m atched control group. In the low-dose subjects (< 31 ppm , m e dian 8-hour TW A of 13.6 ppm [43.4 m g/m 3 ]), the ALC, RBCs, and platelet count were reduced as com p ared with controls. Sim ilarly, in a selected subgroup exposed to a m e dian 8-hour TW A of 7.6 ppm (23 m g /m 3 ) benzene, a statistically significant difference in ALC versus controls was observed ( p <0.03). Furtherm ore, a dose-response relationship was established between the ALC and benzene exposure, as m onitored by organic vapor passive dosim etry and the level of benzene

m e tabolites in the urine. The m e dian 8-hour TW A of 7.6 ppm was designated the LOAEL for these effects and was used to calculate a chronic inhalation RfC for benzene hem a totoxicity in hum ans for com p arison with the RfC derived in the prim a ry analysis using BMD m odeling of the ALC exposure-response data.

135BMD m odeling of the ALC exposure-response data was conducted using the continuous

m odels from EPA’s Benchm ark Dose Modeling Soft ware (version 1.20). In order to fit the data with one of the available m odels, the data were first transform ed according to the equation d 

### 1 = ln(d + 1). The exposure-response data were then fit using the continuous linear m odel, which provided a good fit. The two-degree polynom ial m odel and the power m odel also provided adequate fits to the data; however, the linear m odel was selected because it was m o re parsim onious and the additional param e ters 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 m ean 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 approxim a tely 10% for the proportion of individuals below the 2 nd percentile (or above the 98 th percentile) of the control distribution for norm ally distributed effects. Transform ing the resulting BMC and BMCL back to the original exposure scale yielded a BMC of 13.7 ppm (8-hour TW A) and a BMCL of 7.2
ppm (8-hour TW A).

As suggested in the draft BMD technical guidance docum ent, the BMCL was chosen as the point of departure for the RfC derivation. After converting to m g /m 3 and adjusting for continuous exposure, a BMCL ADJ of 8.2 m g /m 3 was obtained. Dividing this value by an overall UF of 300 yields a chronic inhalation RfC of 3 × 10 -2 mg/m 3 , based on BMD m odeling of the ALC data from the Rothm an et al. (1996a) hum an 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 guida nce for applying an effect-level extrapolation factor to a BMD. In the interim
, a factor of 3 was used in this analysis. For a m o re serious effect, a larger factor, such as 10, m i ght 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/developm ental toxicity study for benzene, com p rise the rem a inder of the 300 com posite UF. These UFs, as well as the UFs for the com p arison analyses and the RfD calculations, are sum m a rized in Table 13.

For com p arison, an RfC was also calculated from the LOAEL of 7.6 ppm (8-hour TW A) from Rothm an et al. (1996a). After converting to m g /m 3 and adjusting for continuous exposure, a LOAEL ADJ of 8.7 m g /m 3 was obtained. Dividing this value by an overall UF of 1000 yields an RfC of 9 × 10 -3 mg / m 3

. 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 (hum an 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/ developm ental toxicity study for benzene. This result of 9 × 10 -3 mg / m 3 based on the LOAEL is in good agreem ent with the result of 3 × 10 -2 mg / m 3 based on the BMCL. The BMD m odeling approach is chosen as the prim a ry analysis in this docum ent because BMD m odeling is a generally superior m e thodology that addresses som e of the lim itations of the LOAEL/NOAEL approach. For exam ple, BMD m odeling m a kes 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 m o re

136Table 13.Summary of uncertainty factors used for deriving the RfC and RfDaBasisBMCL orBMDL(NOAEL orLOAEL)bEffect levelextrapolationfactor(LOAEL-toNOAEL UF)InterspeciesUFIntraspeciesUFSubchronic-to-chronic UFDatabasedeficienciesUFCompositeUFRfD or RfCRfC based onRothman etal. (1996a)(human)8.2 mg/m3(8.7 LOAEL)3(10)NA1033300(1000)0.03 mg/m3(0.009)ComparisonRfC based onWard et al.(1985)(rodent)48.5 mg/m3(17 NOAEL)3(NA)310331000(300)0.05 mg/m3(0.06)RfD based onRothman etal. (1996a)(human)1.2mg/kg/dayc(1.2 LOAEL)3(10)NA1033300(1000)4 × 10-3mg/kg/day(1 × 10-3)ComparisonRfD based onNTP (1986)(rodent)0.7mg/kg/day(18 LOAEL)3(10)1010NA31000(3000)7 × 10-4mg/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 fromthe 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 inhalationabsorption was 50% and oral absorption was 100% in the dose range near the BMC.NA = not applicable.

137standardized point of com p arison across endpoints and studies because it corresponds to a specific response level rather than being m e rely one of the doses used in a particular study.

Support for this chronic inhalation RfC was provided by the experim e ntal anim al study of W a rd et al. (1985). The subchronic inhalation study of W a rd et al. (1985) was selected as a supporting study because it used a relatively long inhalation exposure duration and an adequate

num ber of anim als; it also provided dose-response data. The study identified both a LOAEL of 300 ppm and a NOAEL of 30 ppm

. W a rd et al. exposed m a le and fem a le CD-1 m i ce and Sprague-Dawley rats to 0, 1, 10, 30 or 300 ppm (0, 3.2, 32, 96 or 960 m g /m 3 ) benzene 6 hours/day, 5 days/week for 91 days and m easured various hem a tologic endpoints. The m a le

m ouse appears to be the m o st sensitive sex/species in this study. The exposure-response relationships for the different hem a tologic endpoints in m a le m i ce were m odeled using a BMD

m odeling approach, and decreased HCT was chosen as the critical effect.

BMD m odeling was conducted using the continuous m odels from EPA’s Benchm ark Dose Modeling Software (version 1.20). The linear, polynom ial, and power m odels all resulted in the sam e BMC and BMCL estim ates; however, the linear m odel had better results for the fit statistics. Thus, the continuous linear m odel 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 m ean 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 som e uncertainty about the actual shape of the doseresponse curve in the region of the BMC and, thus, som e corresponding uncertainty about the values of the BMC and BMCL estim ates.

ALCs were not reported in W a rd et al. (1985), so this endpoint could not be com p ared with the hum an ALC results. Total W BC 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, W BC counts did not yield the lowest BMC estim ate. The actual lowest BMC estim ates were obtained for increased m ean cell Hgb (78 ppm ; BMCL = 67
ppm ) and increased m ean cell volum e (79 ppm ; BMCL = 68 ppm ); however, these endpoints are probably not adverse per se. On the other hand, they are likely to be com p ensatory effects—and thus m a rkers of toxicity—and one could probably justify using them as the critical effects. In any event, the BMC estim ates are not m u ch different from the BMC of 100 ppm obtained for decreased HCT. The results are also sim ilar 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 m odel fits were not adequate for the RBC data. Thus, the RBC results have m o re uncertainty.

As suggested in the draft BMD technical guidance docum ent, the BMCL was chosen as the point of departure for the RfC derivation. After converting to m g /m 3 and adjusting for continuous exposure, a BMCL ADJ of 48.5 m g /m 3 was obtained for the critical effect of decreased HCT. To derive the RfC, several UFs are applied to the BMCL ADJ . As discussed above, a UF of 3 is used as an effect-level extrapolation factor, analogous to a LOAEL-to-NOAEL UF, because

138the 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/developm ental 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 hem a tologic fluctuations such as reductions in RBCs and W BCs in the highdose m i ce were noted at interim sacrifice (14 days) as well as at term ination (91 days), suggesting that the responses occurred early in the exposure cycle and then rem a ined

com p aratively unchanged. Dividing the BMCL ADJ by the overall UF of 1000 yields a chronic inhalation RfC of 5 × 10 -2 mg / m 3

. This value is in good agreem ent with the RfC of 3 × 10 -2
mg / m 3 , based on BMD m odeling of the ALC data from the Rothm an et al. (1996a) hum an study.

Sim ilarly, for com p arison purposes, a chronic inhalation RfC can be derived from the NOAEL of 30 ppm observed for hem a tologic effects in the W a rd et al. (1985) experim e ntal anim al study. First, the NOAEL is converted to m g /m 3 and adjusted to equivalent continuous exposure, yielding 17.1 m g /m 3

. UFs are identical to those em ployed 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 / m 3

. This value is also in good agreem ent with the RfC derived from the Rothm an et al. (1996a) study.

For derivation of an RfD, the hum an occupational inhalation study of Rothm an 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 m i ce was selected as a supporting experim e ntal anim al study. These studies provided the best dose-response data for quantitatively evaluating the potential hum an health risks due to oral benzene exposure. The cross-sectional study of Rothm an et al. (1996a) is designated as the principal study because it is a hum an exposure study, with well-quantified exposure estim ates and data on sensitive hem a tological endpoints.

As with the RfC, the BMCL of 7.2 ppm (8-hour TW A) for the default BMR of one standard deviation change from the control m ean response for the critical effect of reduced ALC in Rothm an 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 BMCL ADJ -equivalent oral dose rate of 1.2

m g /kg/day was obtained. The route-to-route extrapolation assum e s 50% absorption of inhaled doses and 100% absorption of oral doses. These assum p tions were based on experim e ntal data (Nom iyam a and Nom i yam a , 1974; Pekari et al., 1992; Srbova et al., 1950; Sabourin et al., 1987) and are the sam e as those used by EPA for route-to-route extrapolation of the inhalation cancer risk estim ates to obtain cancer risk estim ates for oral exposures (U.S. EPA, 1999a). Dividing the BMCL ADJ -oral of 1.2 m g /kg/day by an overall UF of 300 yields a chronic oral RfD of 4 × 10 -3 mg/kg/day , based on BMD m odeling of the ALC data from the Rothm an et al. (1996a) hum an 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/developm ental toxicity study for benzene com p rise the rem a inder of the 300 com posite UF.

139For com p arison, an RfD was also calculated from the LOAEL of 7.6 ppm (8-hour TW A) from Rothm an et al. (1996a). After unit conversion, correction for continuous exposure, routeto-route extrapolation, and division by a com b ined UF of 1000, an RfD of 1 × 10 -3 m g /kg/day was derived from the LOAEL. The route-to-route extrapolation was based on an assum p tion of 50% absorption of inhaled doses and 100% absorption of oral doses, as above. The com b ined UF of 1000 was based on a factor of 10 to account for using a LOAEL because of the lack of anappropriate 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

m g /kg/day is in good agreem ent with the value of 4 × 10 -3 m g /kg/day calculated from the BMC.

Support for this chronic oral RfD was provided by the experim e ntal anim al study of the NTP (1986). In this study, F344 rats and B6C3F1 m i ce of both sexes were adm i nistered benzene by gavage 5 days/week for 103 weeks (see also Section 4.2.1.2). For rats, m a les (50/group) were adm i nistered doses of 0, 50, 100, or 200 m g /kg, and fem a les (50/group) were adm i nistered doses of 0, 25, 50, or 100 m g /kg. B6C3F1 m i ce (50/sex/group) were adm i nistered doses of 0, 25, 50, or 100 m g /kg. Blood was drawn from 10 random ly preselected anim als per species/sex/dose group at 12, 15, 18, and 21 m onths and from all anim als at the term inal kill at 24 m onths. Additional groups of 10 anim als of each sex and species were adm i nistered benzene for 51 weeks at the sam e doses of the 2-year study, and blood was drawn at 0, 3, 6, 9, and 12 m onths. This study identified a LOAEL of 25 m g /kg for leukopenia and lym phocytopenia in fem a le F344 rats and m a le and fem a le B6C3F1 m i ce and 50 m g /kg in m a le F344 rats. These were the lowest doses tested and, thus, no NOAEL was identified.

Reduction in lym phocyte count was selected as the critical effect, and attem p ts were

m a de to m odel the dose-response relationships using a BMD m odeling approach. The m a les of both species exhibited m o re dram a tic and consistent reductions in lym phocyte count than did the fem a les, but it was not clear a priori which species was m o re sensitive; therefore, dose-response analyses were perform ed on datasets for various tim e points for both the m a le m ouse (NTP Table N8) and the m a le rat (NTP Table N4). Various continuous m odels were used to fit the different datasets, as appropriate, and one standard deviation change from the control m e an was used as the BMR, as discussed above. The m odeling results suggested that the m a le rat is m o re sensitive than the m a le m ouse to lym phocyte count reductions from exposure to benzene in this NTP gavage bioassay. However, all but one of the calculated BMDs for the m a le rat were over an order of m a gnitude below the lowest exposure dose of 50 m g /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 im pacts of m odel selection and the uncertainties inherent in extrapolating to lower doses. Nonetheless, BMD m odeling of two subc hronic drinking water studies (Hsieh et al., 1988b; W h ite et al., 1984) supported the selection of a BMD in the range of BMDs estim ated from the NTP (1986) rat data, and a BMDL of 1 m g /kg was selected as the point of departure.

Adjusting the BMDL of 1 m g /kg for exposure 7 days/week yielded a BMDL ADJ of 0.7

m g /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/developm ental study were applied, resulting in a com posite UF of 1000. Dividing the BMDL ADJ by 1000 yields an RfD of 7 × 10 -4 m g /kg/day. This RfD value is in reasonably good agreem ent (within an order of

140m a gnitude) with the RfD value of 4 × 10 -3 m g /kg/day derived from the Rothm an et al. (1996a) hum an inhalation study.

For com p arison purposes, a chronic oral RfD was also calculated using the LOAEL from the NTP (1986) study. The LOAEL of 25 m g /kg was adjusted to a continuous exposure level of 17.9 m g /kg/day and then divided by a UF of 3000 to derive an RfD of 6 × 10 -3 m g /kg/day. The

com b ined 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 (hum an variability), and a factor of 3 for database deficiencies, as above. This value of 6 × 10 -3

m g /kg/day is in good agreem ent with the value 4 × 10

 3 m g /kg/day derived from the Rothm an et al. (1996a) hum an study.

In sum m a ry, the chronic inhalation RfC values calculated on the basis of the hum an and the experim e ntal anim al data and using BMD m odeling and NOAEL/LOAEL approaches are in good agreem ent, yielding values that range from 9 × 10 -3 to 6 × 10 -2 mg / m 3 (Table 14). This consistency in results provides increased confidence in the selected chronic inhalation RfC of 3 × 10 -2 mg / m 3 , which is based on BMD m odeling of the Rothm an et al. (1996a) hum an data.

Sim ilarly, the chronic oral RfD values calculated on the basis of the hum an and the experim e ntal anim al data and using BMD m odeling and NOAEL/LOAEL approaches are in generally good agreem ent, yielding values that range from 7 × 10 -4 to 6 × 10 -3 m g /kg/day. This consistency in results provides increased confidence in the selected chronic oral RfD of 4 × 10 -3

m g /kg/day, which is based on BMD m odeling of the Rothm an et al. (1996a) hum an data.

The RfC and RfD values based on the Rothm an et al. (1996a) hum an study were selected over the values derived from experim e ntal anim al studies because they are based on goodquality hum an data and therefore are not subject to the uncertainties inherent in interspecies extrapolation. Reference values based on BMD m odeling were selected over those calculated using the NOAEL/LOAEL approach because BMD m odeling is a superior m e thodology that

m a kes better use of the exposure-response data. In any event, especially in the case of the RfC, all the estim ates are in sufficiently good agreem ent as to be effectively indistinguishable. The overall confidence in this RfC and RfD assessm ent is m e dium

.

6. MAJOR CONCLUSIONS IN THE CHARACTERIZ ATION OF HAZ ARD AND DOSE RESPONSE

6.1. HUMAN HAZ ARD POTENTIAL

Benzene is widely used as an industrial solvent, as an interm ediate in chem ical syntheses, and as a gasoline additive (NTP, 1994). Because of its widespread use, the potential for hum an exposure is great. The toxicity of benzene has been recognized for m o re than a century, and the

141Table 14. Summary of RfC and RfD estimates using human and experimental animal data, as w e ll as BMD modeling and LOAEL/NOAEL approaches a

Approach RfC (mg/m 3 ) R fD (mg/kg/day)

Human Rodent Human Rodent

BMD m odeling 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 im pacts of benzene exposure have been extensively studied in hum ans and in experim e ntal anim al m odels. Metabolism of benzene is necessary for the com pound’s toxic effects to develop. Evidence has accum u lated indicating that oxidation of benzene by CYP2E1 in the liver is the first step in initiation of benzene toxicity. Convincing evidence of the im portance of CYP2E1 was provided by Valentine et al. (1996), who showed that genetic knockout m i ce lacking expression of the CYP2E1 protein produced m u ch lower levels of benzene m e tabolites and failed to develop signs of genotoxicity and hem a totoxicity following acute benzene exposure at dose levels that resulted in severe genotoxicity and cytotoxicity in both wild-type and B6C3F1 m i ce.

The m a jority of benzene m e tabolism occurs in the liver, but the bone m a rrow is the target organ where its toxicity is expressed with the greatest sensitivity. The m a jor hepatic m e tabolites of benzene are phenol, catechol, and hydroquinone. Catechol and hydroquinone have been shown to accum u late in bone m a rrow after benzene exposure (Rickert et al., 1979). The bone

m a rrow has high peroxidase activity, which results in oxidation of the phenolic m e tabolites produced in the liver to the highly reactive 1,4-benzoquinone (Sm ith et al., 1989). Other target tissues are also characterized by high peroxidase activity (Low et al., 1995). The m e tabolic basis for the toxicity of benzene has been extensively studied (Snyder and Hedli, 1996).

The m o st frequently observed toxic effect of benzene, both in hum ans and test anim al

m odels, is bone m a rrow depression, which leads to lym phocytopenia, leukocytopenia, throm bocytopenia, anem ia, and aplastic anem ia (Aksoy, 1991; Goldstein, 1988; Dosem eci et al., 1996). The m o st sensitive effect observed in hum ans is the depression of ALC in peripheral blood (Rothm an et al., 1996a). In test anim al studies, the m o st sensitive effects observed are depressions of the colony-form ing ability of bone m a rrow 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 hem a topoiesis is a dynam i c process in which stem and progenitor cells, in conjuction with bone m a rrow strom a, give rise to m a ture blood cells. The survival and proliferation of these cells are controlled by m u ltiple growth factors or cytokines that regulate hem a topoiesis. Several studies in m i ce and hum an bone m a rrow cultures have shown that benzene alters cytokine production or response to cytokines (Dem pster and Snyder, 1990; Farris et al., 1993; Cronkite et al., 1989; Irons et al., 1992; Irons and Stillm an, 1993, 1996; Rothm an et al., 1996b)

142Although a large num ber of hum an and experim e ntal anim al studies have been conducted, there are few hum an studies with reliable estim ates of exposure to benzene and few long-term
, repeated-dose experim e nts in test anim als. Hum a n studies frequently are also

com p licated by exposure to other solvents. The long-term test anim al 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 exam ined. Thus, the lack of reliable NOAEL values from either hum an or test anim al studies is an area of uncertainty in establishing RfD and RfC values that are protective of hum an health. The use of BMD

m odeling 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 estim ate BMDs.

Another area of scientific uncertainty in this assessm ent concerns the neurotoxic effects of benzene. As is the case with m a ny other organic solvents, benzene has been shown to produce neurotoxic effects in experim e ntal anim als and hum ans after short-term exposures to relatively high concentrations of the com pound. Benzene produces generalized sym p tom s such as dizziness, headache, and vertigo, leading to drowsiness, trem or, delirium
, and loss of consciousness. In an occupational study, Kahn and Muzyka (1973) reported that workers

com p lained of frequent headaches (usually at the end of the workday), tired easily, had difficulties sleeping, and com p lained of m e m o ry loss. Overall, there is a lack of reliable inform ation on dose-related neurotoxic effects under low-dose chronic exposure conditions in either hum ans or experim e ntal anim al m odel system s. Li et al. (1992) reported that forelim b grip strength and the frequency of rapid response in Y-m aze running in Kunm ing m i ce was increased following brief inhalation exposure to 0.78 ppm (2.5 m g /m 3 ) benzene, but both responses were decreased at concentrations of 3.13 ppm (10 m g /m 3 ) or higher. Several experim e ntal deficiencies in this study prevent its use for calculating the risk to hum ans, but further investigation of these effects could reveal neurotoxic effects of concern to hum an health.

There have been a num ber of developm ental and reproductive studies in hum ans and in test anim al m odel system s. Several test anim al studies have shown developm ental effects exhibiting m a nifestations such as reductions in num bers of live fetuses, reductions in live weight, and m i nor skeletal variants, but the benzene concentrations used also caused severe

m a ternal toxicity. The studies by Keller and Snyder (1986, 1988), however, dem onstrated that exposure to low concentrations of 5, 10, or 20 ppm (16, 32, or 64 m g /m 3 ) benzene in utero during developm ent caused changes in colony-form ing hem a topoietic cells. These studies were considered supporting, because the LOAEL of 5 ppm is below the LOAEL of 7.6 ppm established for hem a totoxic and im m unotoxic effects observed in hum ans (Rothm an et al., 1996a). However, the confidence in the hum an data is m u ch higher than that in the lim ited hem a totoxic endpoints m easured in m i ce by Keller and Snyder (1986, 1988). Furtherm ore, the responses did not establish consistent patterns in different ages of the progeny. Also, a lim ited

num ber of anim als were exam ined. These lim itations m a ke this study less useful because of the high degree of uncertainty that these are truly adverse effects. (See section 4.2.2.2.)

1436.2. DOSE RESPONSE

6.2.1. Inhalation RfC

Quantitative estim ates of hum an health risk as a result of low-level chronic exposure to benzene via the inhalation route are based on data from hum an occupational inhalation exposure and from a subchronic inhalation study in experim e ntal anim als. Hem a totoxicity and im m unotoxicity are the critical effects observed in both hum ans and test anim als.

The air concentration of benzene considered to be without any appreciable risk with lifetim e chronic hum an exposure (the RfC) is 3 × 10 -2 mg / m 3

. This value was obtained by applying a UF of 300 to the BMCL ADJ obtained from BMR m odeling of the exposure-response data for reductions in ALC from the Rothm an et al. (1996a) occupational study, using a BMR level of one standard deviation change from the control m ean response. This RfC is in good
agreem ent with the value of 9 × 10 -3 mg / m 3 based on the LOAEL in the Rothm an et al. (1996a) study, with a UF of 1000.

The RfC of 3 × 10 -2 mg / m 3 is sim ilarly in good agreem ent with the value of 5 × 10 -2
mg / m 3 obtained by applying a UF of 1000 to the BMCL ADJ estim ated from BMD m odeling of the exposure-response data for the hem a tologic effect of decreased HCT in m a le m i ce from the W a rd et al. (1985) subchronic inhalation study, us ing a BMR level of one standard deviation change from the control m ean response. Finally, both of these values are in good agreem ent with the value of 6 × 10 -2 mg / m 3 based on the NOAEL in the W a rd et al. (1985) study, with a UF of 300.

The overall confidence in this RfC assessm ent is m e dium

. The Rothm an et al. (1996a) and W a rd et al. (1985) studies were both well conducted, and various m e thodologies for deriving the RfC yielded sim ilar results for the two studies. Furtherm ore, the availability of good-quality hum an data for a sensitive endpoint elim inates the uncertainty associated with basing the RfC on experim e ntal anim al data. However, with continuous endpoints such as the hem a tologic
param e ters m easured in these studies, there is uncertainty about when a change in a param e ter that has inherent variability becom e s an adverse effect. Other uncertainties explicitly recognized in the quantitative derivation include intraspecies variability, that is, the need to accom m odate sensitive hum an subgroups; the extrapolation of subchronic results to a lifetim e exposure scenario; and database deficiencies.

6.2.2. Oral RfD

Quantitative estim ates of hum an health risk as a result of low-level chronic exposure to benzene via the oral route are based on data from hum an occupational inhalation exposure and from a chronic gavage study in experim e ntal anim als. Once again, hem a totoxic responses are the critical effects observed.

The hum an chronic dose of ingested benzene considered to be without any appreciable risk (the RfD) is 4 × 10 -3 m g /kg/day. This value was obtained by applying a UF of 300 to the oral equivalent dose extrapolated from the BMCL ADJ obtained from BMD m odeling of the ALC data from the hum an occupational study of Rothm an et al. (1996a) using a BMR level of one

144standard deviation change from the control m ean response, as for the RfC discussed above. This RfD is in good agreem ent with the value of 1 × 10 -3 m g /kg/day based on the oral equivalent LOAEL from the Rothm an et al. (1996a) study, with a UF of 1000.

This RfD of 4 × 10 -3 m g /kg/day is also in good agreem ent with the values of 7 × 10 -4

m g /kg/day and 6 × 10 -3 m g /kg/day, which are 1/1000 of the BMDL ADJ derived from the m a le 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 assessm ent is m e dium

. The Rothm an et al. (1996a) study was well conducted, and the availability of good-quality hum an data for a sensitive endpoint elim inates the uncertainty associated with basing the RfD on experim e ntal anim al data. However, with continuous endpoints such as hem a tologic param e ters, there is uncertainty about when a change in a param e ter that has inherent variability becom e s an adverse effect. Other uncertainties explicitly recognized in the quantitative derivation include intraspecies variability,
to accom m odate sensitive hum an subgroups; the extrapolation of subchronic results to a lifetim e exposure scenario; and database deficiencies.

A further uncertainty in the RfD ensues from the use of route-to-route extrapolation to estim ate oral equivalent doses from inhalation exposures resulting from analysis of the Rothm an et al. (1996a) occupational data. In experim e nts conducted to com p are the m e tabolite doses to the target organ following oral or inhalation exposure, Sabourin et al. (1987, 1989) found that there was no sim p le relationship between the two routes of exposure. Oral doses and inhalation exposures that produced sim ilar concentrations of one m e tabolite in the blood produced very different doses of another m e tabolite. The target specificity of benzene toxicity for the bone

m a rrow progenitor cells irrespective of route of adm i nistration, however, is well docum ented both in hum ans and experim e ntal anim al m odels. Thus, route-to-route extrapolation is justified and introduces a lower degree of uncertainty than extrapolating from test anim als to hum ans (U.S. EPA, 1999a). Use of a m odifying factor of3 to account for uncertainty in the route-toroute extrapolation was considered; however, it was deem ed unnecessary. The RfD is based on hum an data for a sensitive endpoint; thus, it was felt that the com posite UF of 300 provides sufficient protection.

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