Benzene (CASRN 71-43-2) 

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0276

Benzene; CASRN 71-43-2; 04/17/2003

Health assessment information on a chemical substance is included in
IRIS only after a comprehensive review of chronic toxicity data by U.S.
EPA health scientists from several Program Offices and the Office of
Research and Development. The summaries presented in Sections I and II
represent a consensus reached in the review process. Background
information and explanations of the methods used to derive the values
given in IRIS are provided in the Background Documents. 

STATUS OF DATA FOR Benzene

File First On-Line 03/01/1988

Category (section)	Status	Last Revised

Oral RfD Assessment (I.A.)	on-line	04/17/2003 

Inhalation RfC Assessment (I.B.)	on-line	04/17/2003

Carcinogenicity Assessment (II.)	on-line 	01/19/2000 

_I.  Chronic Health Hazard Assessments for Noncarcinogenic Effects

_I.A. Reference Dose for Chronic Oral Exposure (RfD)

Substance Name — Benzene

CASRN — 71-43-2

Last Revised — 04/17/2003 

The oral Reference Dose (RfD) is based on the assumption that thresholds
exist for certain toxic effects such as cellular necrosis. It is
expressed in units of mg/kg/day. In general, the RfD is an estimate
(with uncertainty spanning perhaps an order of magnitude) of a daily
exposure to the human population (including sensitive subgroups) that is
likely to be without an appreciable risk of deleterious effects during a
lifetime. Please refer to the IRIS Background Document for an
elaboration of these concepts. The U.S. EPA has evaluated this substance
for potential human carcinogenicity. A summary of that evaluation is
found in Section II of this file. 

__I.A.1. Oral RfD Summary

Critical Effect	Experimental Doses*	UF	MF	RfD

Decreased lymphocyte 

count (Human occupational

inhalation study;

Rothman et al., 1996)	BMDL = 1.2 mg/kg/day	300	1	4.0 x 10-3

mg/kg/day

*Conversion factors: MW = 78.11. Assuming 25ºC and 760 mm Hg, BMCL
(mg/m3) = 7.2 ppm x MW/24.45 = 23 mg/m3. BMCLADJ = 23 mg/m3 x 10 m3/20
m3 x 5 days/7days = 8.2 mg/m3. The BMDL was derived by route-to-route
extrapolation with the assumptions that inhalation absorption was 50%
and oral absorption was 100% in the dose range near the BMC. BMDLADJ =
8.2 mg/m3 x 20 m3/day x 0.5 ÷ 70 kg = 1.2 mg/kg/day. (The original BMC
was based on a benchmark response of one standard deviation change from
the control mean.)

__I.A.2. Principal and Supporting Studies (Oral RfD)

The RfD is based on route-to-route extrapolation of the results of
benchmark dose (BMD) modeling of the absolute lymphocyte count (ALC)
data from the occupational epidemiologic study by Rothman et al. (1996),
in which workers were exposed to benzene by inhalation. A comparison
analysis based on BMD modeling of data from the National Toxicology
Program's (NTP's) experimental animal gavage study (NTP, 1986) was also
conducted. In addition, comparison analyses using the
lowest-observed-adverse-effect levels (LOAELs) from the Rothman et al.
(1996) and NTP (1986) studies were performed.

Rothman et al. (1996) conducted a cross-sectional study of 44 workers
exposed to benzene and 44 age- and gender-matched unexposed controls.
Twenty-one of the 44 subjects in the exposed and control groups were
female. Mean (standard deviation) years of occupational exposure to
benzene were 6.3 (4.4), with a range of 0.7-16 years. Benzene exposure
was monitored by organic vapor passive dosimetry badges worn by each
worker for a full workshift on 5 days within a 1-2 week period prior to
collection of blood samples. The median 8-hour time-weighted average
(TWA) benzene exposure concentration for all exposed workers was 31 ppm
(99 mg/m3). The exposed group was subdivided into two equal groups of 22
subjects: those exposed to greater than the median concentration and
those exposed to less than the median concentration. The median 8-hour
TWA exposure concentration was 13.6 ppm (43.4 mg/m3) for the
low-exposure group and 91.9 ppm (294 mg/m3) for the high-exposure group.


Six hematological measurements were evaluated: total white blood cell
(WBC) count, ALC, hematocrit, red blood cell (RBC) count, platelet
count, and mean corpuscular volume (MCV). All six parameters were
significantly different in the high-benzene exposure group (>31 ppm)
when compared to controls. ALC, WBC count, RBC count, hematocrit, and
platelets were all significantly decreased, and MCV was significantly
increased. ALC was the most sensitive endpoint; it was reduced from 1.9
x 103/µL blood in controls to 1.6 x 103/µL (p<0.01) in the <31 ppm
group and to 1.3 x 103/µL (p<0.001) in the group exposed to >31 ppm
benzene. The ALC was also significantly reduced (1.6 x 103/µL; p=0.03)
in a subgroup of 11 workers exposed to a median 8-hour TWA of 7.6 ppm
(24 mg/m3) benzene. For additional details about this study see Section
I.B.2. 

BMD modeling of the ALC data of Rothman et al. (1996) yielded a
benchmark concentration (BMC) of 13.7 ppm (8-hr TWA) and a BMCL (the 95%
lower bound on the BMC) of 7.2 ppm (8-hr TWA) for the default benchmark
response of one standard deviation change from the control mean (see
Section I.B.2 for details of the analysis). Converting the units and
adjusting for continuous exposure results in a BMCLADJ of 8.2 mg/m3.
[According to the Ideal Gas Law, concentration in mg/m3 = concentration
in ppm x MW/24.45 at 25ºC and 760 mm Hg. Thus, BMCL (mg/m3) = 7.2 x
78.11/24.45 = 23.0 mg/m3. BMCLADJ = 23.0 mg/m3 x 10 m3/20 m3 x 5 days/7
days = 8.2 mg/m3, where 10 m3 is the default human occupational volume
of air inhaled in an 8-hour workshift, and 20 m3 is the default human
ambient volume of air inhaled in a 24-hour day (U.S. EPA, 1994).]

In the support document for the benzene cancer assessment on IRIS (U.S.
EPA, 1999), EPA provided a simple method for extrapolation of
benzene-induced cancer risk from the inhalation to the oral route. The
same method is applied here for noncancer (hematopoietic) effects. The
method is based on the relative efficiency of benzene absorption across
routes of exposure, especially pulmonary and gastrointestinal barriers.
An inhalation absorption rate of 50% and an oral absorption rate of 100%
were used to calculate the absorbed benzene dose. These values are based
on human inhalation absorption studies and the study by Sabourin et al.
(1987) that compared inhalation and oral absorption in rats and mice.
The authors found that during a 6-hour inhalation exposure, the
retention of [14C]benzene decreased from 33 ± 6% to 15 ± 9% for rats
and from 50 ± 1% to 10 ± 2% for mice as exposure concentration
increased from 26 to 2,600 mg/m3 (10 to 1,000 ppm). In the same study,
gastrointestinal absorption of benzene administered by gavage was >97%
for doses between 0.5 and 150 mg/kg body weight. At oral doses below 15
mg/kg, >90% of the 14C excreted was in the urine as non-ethyl
acetate-extractable material. At higher doses, an increasing percentage
of the orally administered benzene was exhaled unmetabolized. Thus, in
the dose range represented by the BMCL from the study by Rothman et al.
(1996), absorption of a comparable oral dose was assumed to be 100%. See
also U.S. EPA (1999) for more details about the route-to-route
extrapolation of benzene inhalation results to oral exposures.

To calculate an equivalent oral dose rate, the BMCLADJ is multiplied by
the default inhalation rate, multiplied by 0.5 to correct for the higher
oral absorption, and divided by the standard default human body weight
of 70 kg: 8.2 mg/m3 x 20 m3/day x 0.5 ÷ 70 kg = 1.2 mg/kg/day. The RfD
is then derived by dividing the equivalent oral dose by the overall
uncertainty factor (UF) of 300: RfD = equivalent oral dose/UF = 1.2
mg/kg/day ÷ 300 = 4 x 10-3 mg/kg/day. The overall UF of 300 comprises a
UF of 3 for effect-level extrapolation, 10 for intraspecies differences
(human variability), 3 for subchronic-to-chronic extrapolation, and 3
for database deficiencies (see Section I.A.3).

For comparison, an RfD was also calculated based on the LOAEL of 7.6 ppm
(8 hr TWA) from the Rothman et al. (1996) study (see Section I.B.2).
Converting the units and adjusting for continuous exposure results in a
LOAELADJ of 8.7 mg/m3. Then the equivalent oral exposure is calculated
as above: 8.7 mg/m3 x 20 m3/day x 0.5 ÷ 70 kg = 1.2 mg/kg/day. The
equivalent oral exposure is then divided by an overall UF of 1000 to
obtain the RfD: 1.2 mg/kg/day ÷ 1000 = 1 x 10-3 mg/kg/day. The combined
UF of 1000 represents UFs of 10 to account for the use of a LOAEL
because of the lack of an appropriate no-observed-adverse-effect level
(NOAEL), 10 for intraspecies differences in response (human
variability), 3 for subchronic-to-chronic extrapolation, and 3 for
database deficiencies. The value of 1 x 10-3 mg/kg/day is in good
agreement with the value of 4 x 10-3 mg/kg/day calculated from the BMDL
(the 95% lower bound on the BMD).

A comparison RfD derivation was also performed using the results of the
NTP (1986) experimental animal gavage study. In that study, F344 rats
and B6C3F1 mice of both sexes were administered benzene by gavage, 5
days/week for 103 weeks. Male rats (50/group) were administered doses of
0, 50, 100, or 200 mg/kg, and females (50/group) were administered doses
of 0, 25, 50, or 100 mg/kg. B6C3F1 mice (50/sex/group) were administered
doses of 0, 25, 50, or 100 mg/kg. Blood was drawn from 10 randomly
preselected animals per species/sex/dose group at 12, 15, 18, and 21
months, as well as from all animals at the terminal kill at 24 months.
Additional groups of 10 animals of each sex and species were
administered benzene for 51 weeks at the same doses of the 103-week
(2-year) study, and blood was drawn at 0, 3, 6, 9, and 12 months. This
study identified a LOAEL of 25 mg/kg for leukopenia and lymphocytopenia
in female F344 rats and male and female B6C3F1 mice and 50 mg/kg in male
F344 rats. These were the lowest doses tested, and thus no NOAEL was
identified.

Reductions in lymphocyte count was the critical effect, and attempts
were made to model the dose-response relationships using a BMD modeling
approach. Modeling was performed for each dataset in two data groupings
within which the datasets are comparable (6- and 9-month; and
12-,15-,18-, and 21-month), and ranges of results are presented. Each of
these datasets had at most 10 animals/dose, so the dose-response results
are not very robust. The males of each species exhibited more dramatic
and consistent reductions in lymphocyte count, but it was not clear a
priori which species was more sensitive; therefore, dose-response
analyses were performed for both the male mouse and the male rat. 

The continuous linear, polynomial, and power models in EPA's Benchmark
Dose Modeling Software (version 1.20) were used for the modeling. The
software estimates the parameters using the method of maximum
likelihood. Most of the data were supralinear (i.e., the magnitude of
the reductions in lymphocyte count decreased with increasing unit dose),
and it was necessary to transform the dose data according to the formula
d’ = ln(d+1) in order to fit the available models. The results are
summarized in Table 1. For each dataset, the selected model was chosen
based on the lowest Akaike's Information Criterion (AIC) value, with
consideration of the graphical display, as suggested in EPA's draft
Benchmark Dose Technical Guidance Document (U.S. EPA, 2000). For
selecting between models within a family of models, for example, between
a linear and a two-degree polynomial model, consideration was given to
the log-likelihood values to evaluate the statistical significance of
adding an extra parameter. There was substantial variability in these
data, but it appeared to be random and not amenable to modeling.
Therefore, constant variance was assumed for all the models, although in
some cases the variances failed the test for homogeneity. 

In the absence of a clear definition for an adverse effect for this
endpoint, a default benchmark response of one standard deviation change
from the control mean response was selected, as suggested in the draft
technical guidance document. This definition of the benchmark response
is highly sensitive to the substantial variability in data such as
these, and thus the benchmark response itself is not very robust. The
usefulness of this default definition would be strengthened by the use
of a larger dataset of historical control data, but such data were not
located. The software uses the estimated "constant" standard deviation
as the standard deviation for all the group means. The 95% lower
confidence limits (BMDLs) on the BMDs are calculated using the
likelihood profile method.

The results shown in Table 1 suggest that the male rat is more sensitive
than the male mouse to lymphocyte count reductions from exposure to
benzene in this NTP gavage bioassay because the ranges of BMDs/BMDLs are
substantially lower for the male rat, especially for year 2. The ranges
for the male rat are fairly tight, and the models selected provide good
fits to all the male rat datasets. However, all but one of the
calculated BMDs for the male rat are over an order of magnitude below
the lowest exposure dose of 50 mg/kg. Ideally, BMDs should be closer to
the low end of the range of observation, that is, the range of the
actual exposure doses, to reduce the impacts of model selection and the
uncertainties inherent in extrapolating to lower doses.

Nevertheless, data from two drinking water studies provide support for
selecting a BMD in this range. These two studies were of shorter
duration and used fewer experimental animals than the NTP (1986) study;
however, they do provide dose-response data for BMD modeling, and they
also have the advantage of being drinking water studies; thus the
benzene exposure scenario is more relevant to human oral benzene
exposures. In one study, Hsieh et al. (1988) exposed male CD-1 mice
(five/group) to 0, 8, 40, or 180 mg/kg/day benzene in drinking water for
28 days. Hematological effects were observed at all exposure levels. BMD
modeling of the ALC yielded a BMD of 2.2 mg/kg/day and a BMDL of 1.4
mg/kg/day, based on a linear model with transformed doses and a
benchmark response of one standard deviation change from the control
mean, as above. In the second study, White et al. (1984) exposed female
B6C3F1 mice to 0, 12, 195, or 350 mg/kg/day benzene in drinking water
for 30 days. BMD modeling of the ALC (five to six mice/group) resulted
in a BMD of 11.6 mg/kg/day and a BMDL of 5.3 mg/kg/day (also based on a
linear model with transformed doses and a benchmark response of one
standard deviation change from the control mean, as above).

Table 1. BMD modeling results for NTP (1986) male mouse and male rat
lymphocyte counts, with transformed dose data

Dataset	Model	Variance

Homogeneity	Fit	BMDa

(mg/kg)	BMDLa (mg/kg)

Male Mouse

6-month	two-degree

polynomial	ok	borderline

p=0.047	19.68	6.57

9-month	linear	no	yes, p=0.35	9.07	4.05

year 1 range	9.07-19.68	4.05-6.57

12-month	linear	ok	yes, p=0.30	3.74	2.32

15-month	power	no	yes, p=0.31	47.46	18.55

18-month	power	no	borderline

p=0.09	28.93	13.99

21-month	power	no	yes, p=0.15	23.34	5.80

year 2 range	3.74-47.46	2.32-18.55

Male Rat

6-month	power	ok	yes, p=0.30	9.92	4.52

9-month	linear	no	yes, p=0.11	3.71	2.30

year 1 range	3.71-9.92	2.30-4.52

12-month	linear	no	yes, p=0.22	1.34	0.95

15-month	linear	ok	yes, p=0.93	1.34	0.95

18-month	linear	no	yes, p=0.22	2.73	1.74

21-month	linear	ok	yes, p=0.54	1.69	1.10

year 2 range	1.34-2.73	0.95-1.74

aUnadjusted animal dose in mg/kg, after transforming the results back
according to the formula dose = exp(transformed dose) – 1. (The BMD
was based on a benchmark response of one standard deviation change from
the control mean.)

The results in Table 1 from BMD modeling of the male rat ALC data from
the NTP (1986) study show the lowest BMDL of about 1 mg/kg at three time
points in the second year; thus this was selected as the point of
departure for an RfD calculation. Adjusting for exposure 7 days/week
yields a BMDLADJ of 0.7 mg/kg/day. This value is divided by an overall
UF of 1000 to obtain the RfD: RfD = 0.7 mg/kg/day ÷ 1000 = 7 x 10-4
mg/kg/day. The overall UF of 1000 comprises UFs of 3 for effect-level
extrapolation, 10 for interspecies extrapolation for oral studies, 10
for intraspecies variability, and 3 for database deficiencies. This RfD
value is in reasonably good agreement (within an order of magnitude)
with the RfD of 4 x 10-3 mg/kg/day derived from the Rothman et al.
(1996) human inhalation study.

For comparison purposes, an RfD can also be derived from the LOAEL of 25
mg/kg identified for hematological effects in the NTP (1986) study
(there was no NOAEL). Adjusting from 5-day to 7-day exposure yields a
LOAELADJ of 18 mg/kg/day, which can be used to calculate an RfD for
benzene as follows: RfD = LOAELADJ ÷ UF = 18 mg/kg/day ÷ 3000 = 6 x
10-3 mg/kg/day, where the combined UF of 3000 is made up of component
factors of 10 for LOAEL-to-NOAEL extrapolation, 10 for interspecies
extrapolation, 10 for intraspecies variability, and 3 for database
deficiencies. This value is in good agreement with the RfD of 4 x 10-3
mg/kg/day calculated from the BMD analysis of the Rothman et al. (1996)
human data.

__I.A.3. Uncertainty and Modifying Factors (Oral RfD)

UF = 300 for the BMCL-oral-equivalent from the Rothman et al. (1996)
study.

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, based on the professional judgement that, although the BMD
corresponds to an adverse effect level at the low end of the observable
range, the endpoint is not very serious in and of itself. Decreased ALC
is a very sensitive sentinel effect that can be measured in the blood,
but it is not a frank effect, and there is no evidence that it is
related to any functional impairment at levels of decrement near the
benchmark response. For a more serious effect, a larger factor, such as
10, might be selected. Second, a factor of 10 was used for intraspecies
differences in response (human variability) as a means of protecting
potentially sensitive human subpopulations. Third, a
subchronic-to-chronic extrapolation factor was applied because the mean
exposure duration for the subjects in the principal study was 6.3 years,
which is less than the exposure duration of 7 years (one-tenth of the
assumed human life span of 70 years) that has been used by the Superfund
program as a cut-off for deriving a subchronic human reference dose
(U.S. EPA, 1989). Furthermore, the exposure duration varied from 0.7
years to 16 years. However, because the mean exposure duration was near
the borderline of what would be considered chronic (i.e., 6.3 years vs.
7 years), a 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 no two-generation reproductive and developmental toxicity
studies for benzene are available. Therefore, an overall UF of 3 x 10 x
3 x 3 = 300 is used to calculate the chronic oral RfD.

For the comparison analysis based on the Rothman et al. (1996)
LOAELADJ-equivalent oral dose rate value of 1.2 mg/kg/day, the following
UFs were selected: a factor of 10 for use of a LOAEL due to lack of an
appropriate NOAEL, a factor of 10 for intraspecies variability, a factor
of 3 for subchronic-to-chronic extrapolation, and a factor of 3 for
database deficiencies, as above. Hence, an overall UF of 10 x 10 x 3 x 3
= 1000 was used in the comparison analysis.

For the comparison analysis based on the BMDLADJ calculated from BMD
modeling of the male rat data from the NTP (1986) gavage study, the
following UFs were used: a UF of 3 for effect-level extrapolation, which
is analogous to the LOAEL-to-NOAEL extrapolation factor, because the BMC
is considered an adverse effect level; 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. Thus, an overall UF of 3 x 10 x
10 x 3 = 1000 was used in this comparison analysis. 

Finally, for the comparison analysis based on the LOAEL from the NTP
(1986) gavage study, the following UFs were used: 10 for LOAEL-to-NOAEL
extrapolation, 10 for interspecies extrapolation, 10 for intraspecies
variability, and 3 for database deficiencies. Therefore, an overall UF
of 3000 was used in this comparison analysis.

__I.A.4. Additional Studies/Comments (Oral RfD)

Benzene is toxic by all routes of administration. Hematotoxicity and
immunotoxicity have been consistently reported to be the most sensitive
indicators of noncancer toxicity in both humans and experimental
animals, and these effects have been the subject of several reviews
(Aksoy, 1989; Goldstein, 1988, Snyder et al., 1993; Ross, 1996; U.S.
EPA, 2002). The bone marrow is the target organ for the expression of
benzene hematotoxicity and immunotoxicity. Leukocytopenia has been
consistently shown to be a more sensitive indicator of benzene toxicity
in experimental animal systems than anemia, and lymphocytopenia has been
shown to be an even more sensitive indicator of benzene toxicity than
overall leukocytopenia. Neither gastrointestinal effects from oral
exposure nor pulmonary effects due to inhalation exposure have been
reported. (see Section I.B.4 for a more detailed summary of benzene
toxicity). 

For more detail on Susceptible Populations, exit to   HYPERLINK
"http://www.epa.gov/iris/toxreviews/0276-tr.pdf" \l "page=124"  the
toxicological review, Section 4.4  (PDF). 

__I.A.5. Confidence in the Oral RfD

Study — Medium

Database — Medium

RfD — Medium

The overall confidence in this RfD assessment is medium. The principal
study of Rothman et al. (1996) was well conducted, and the availability
of good-quality human data for a sensitive endpoint eliminates the
uncertainty associated with basing the RfD on experimental animal data.
A dose-response relationship was established between ALC and benzene air
concentration and benzene urine metabolites. Six blood parameters
measured (ALC, WBC count, RBC count, hematocrit, platelets, and MCV)
were significantly different in the high- benzene-exposure group when
compared with controls. However, only the ALC was reduced in a subgroup
of 11 subjects exposed to a median 8-hour TWA of 7.6 ppm benzene,
suggesting that this exposure level may be at the low end of the range
of benzene exposures eliciting hematotoxic effects in humans. 

In addition, the RfD of 4 x 10-3 mg/kg/day obtained from route-to-route
extrapolation of the BMD modeling results from the Rothman et al. (1996)
study is in good agreement with the value of 1 x 10-3 mg/kg/day based on
the oral equivalent LOAEL. The RfD is also in good agreement with the
value of 7 x 10-4 mg/kg/day, based on BMD modeling of the male rat ALC
data from the NTP (1986) chronic rodent gavage study and the value of 6
x 10-3 mg/kg/day based on the LOAEL from the NTP (1986) study. 

With continuous endpoints such as hematological parameters, there is
uncertainty about when a change in a parameter that has inherent
variability becomes an adverse effect. Other uncertainties explicitly
recognized in the quantitative derivation of the chronic oral RfD
include intraspecies variability (to accommodate sensitive human
subgroups), the applicability of the subchronic inhalation data to
chronic oral exposures, and database deficiencies due to the lack of a
two-generation reproductive/developmental toxicity study for benzene. 

Route-to-route extrapolation was used to estimate oral equivalent doses
from inhalation exposures resulting from analysis of the Rothman et al.
(1996) occupational data. In experiments conducted to compare the
metabolite doses to the target organ following oral or inhalation
exposure, Sabourin et al. (1987, 1989) found that there was no simple
relationship between the two routes of exposure. All published
experimental animal models of the in vivo metabolism and disposition of
benzene have used the physiologically based approach to
pharmacokinetics, and they conclude that formation of metabolites follow
Michaelis-Menten kinetics. Although these models predict the urinary
metabolites formed from benzene exposures, they offer no information
regarding the dosimetry of oxidative metabolites in the bone marrow, a
site of action. However, the target specificity of benzene toxicity for
the bone marrow progenitor cells irrespective of route of administration
is well documented in both humans and experimental animal models. Thus,
route-to-route extrapolation is justified and introduces a lower degree
of uncertainty than extrapolating from test animals to humans (U.S. EPA,
1999). Use of a modifying factor of 3 was considered to recognize
uncertainties in the route-to-route extrapolation; however, it was
deemed unnecessary. The RfD is based on human data for a sensitive
endpoint; thus, it was felt that the composite UF of 300 provides
sufficient protection. 

For more detail on Characterization of Hazard and Dose Response, exit to
  HYPERLINK "http://www.epa.gov/iris/toxreviews/0276-tr.pdf" \l
"page=156"  the toxicological review, Section 6  (PDF). 

__I.A.6. EPA Documentation and Review of the Oral RfD

Source Document — U.S. EPA, 2002

This assessment was peer reviewed by external scientists as well as in
response to public comments. Their comments have been evaluated
carefully and incorporated in the finalization of this IRIS Summary. The
  HYPERLINK "http://www.epa.gov/iris/supdocs/benzene_nc-pr.pdf"  peer
review document  (12 pages, 135 Kbytes) is available in Adobe PDF
format.

Other EPA Documentation — U.S. EPA, 1985, 1999

Date of Agency Consensus — January 23, 2002

__I.A.7. EPA Contacts (Oral RfD)

Please contact the IRIS Hotline for all questions concerning this
assessment or IRIS, in general, at (202)566-1676 (phone), (202)566-1749
(FAX) or   HYPERLINK "mailto:hotline.iris@epa.gov"  hotline.iris@epa.gov
 (internet address). 

  HYPERLINK "http://www.epa.gov/iris/subst/0276.htm" \l
"content#content"  Top of page 

_I.B. Reference Concentration for Chronic Inhalation Exposure (RfC)

Substance Name — Benzene

CASRN — 71-43-2

Last Revised — 04/17/2003

The inhalation Reference Concentration (RfC) is analogous to the oral
RfD and is likewise based on the assumption that thresholds exist for
certain toxic effects such as cellular necrosis. The inhalation RfC
considers toxic effects for both the respiratory system
(portal-of-entry) and for effects peripheral to the respiratory system
(extrarespiratory effects). It is generally expressed in units of
mg/cu.m. In general, the RfC is an estimate (with uncertainty spanning
perhaps an order of magnitude) of a daily inhalation exposure of the
human population (including sensitive subgroups) that is likely to be
without an appreciable risk of deleterious effects during a lifetime.
Inhalation RfCs were derived according to Methods for Derivation of
Inhalation Reference Concentrations and Application of Inhalation
Dosimetry (U.S. EPA, 1994). RfCs can also be derived for the
noncarcinogenic health effects of substances that are carcinogens.
Therefore, it is essential to refer to other sources of information
concerning the carcinogenicity of this substance. If the U.S. EPA has
evaluated this substance for potential human carcinogenicity, a summary
of that evaluation will be contained in section II of this file.

__I.B.1. Inhalation RfC Summary

Critical Effect	Exposures*	UF	MF	RfC

Decreased lymphocyte

count (Human occupational

inhalation study of 

Rothman et al., 1996)	BMCL = 8.2 mg/m3	300	1	3 x 10-2

mg/m3

*Conversion factors: MW = 78.11. BMCL = 7.2 ppm, 8-hour TWA. Assuming
25ºC and 760 mm Hg, BMCL (mg/m3) = 7.2 ppm x MW/24.45 = 23.0 mg/m3.
BMCLADJ = 23.0 mg/m3 x 10 m3/20 m3 x 5 days/7days = 8.2 mg/m3. (The BMC
was based on a benchmark response of one standard deviation change from
the control mean.) 

__I.B.2. Principal and Supporting Studies (Inhalation RfC)

The RfC is based on BMD modeling of the ALC data from the occupational
epidemiologic study of Rothman et al. (1996), in which workers were
exposed to benzene by inhalation. A comparison analysis based on BMD
modeling of hematological data from the Ward et al. (1985) subchronic
experimental animal inhalation study was also conducted. In addition,
comparison analyses using the LOAEL from the Rothman et al. (1996) study
and the NOAEL from the Ward et al. (1985) study were performed.

Rothman et al. (1996) conducted a cross-sectional study of 44 workers
exposed to a range of benzene concentrations and 44 age- and
gender-matched unexposed controls, all from Shanghai, China. Twenty-one
of the 44 subjects in the exposed and control groups were female. The
exposed workers were from three workplaces where benzene was used–a
factory that manufactured rubber padding for printing presses, a factory
that manufactured adhesive tape, and a factory that used benzene-based
paint. The unexposed workers were from two workplaces: a factory that
manufactured sewing machines and an administrative facility. Workers who
had a prior history of cancer, therapeutic radiation, chemotherapy, or
current pregnancy were excluded. Requirements for inclusion in the study
were current employment for at least 6 months in a factory that used
benzene, minimal exposure to other aromatic solvents, and no exposure to
other chemicals known to be toxic to bone marrow or to ionizing
radiation. Controls who had no history of occupational exposure to
benzene or other bone marrow-toxic agents were frequency-matched to the
exposed subjects on age (5-year intervals) and gender.

Benzene exposure was monitored by organic vapor passive dosimetry badges
worn by each worker for a full workshift on 5 days within a 1-2 week
period prior to collection of blood samples. Benzene exposure of
controls in the sewing machine factory was monitored for 1 day, but no
exposure monitoring was performed in the administrative facility.
Benzene exposure was also evaluated by analyzing for benzene metabolites
in urine samples collected at the end of the benzene exposure period for
the exposed subjects. Historical benzene exposure of the subjects was
evaluated by examining employment history. Data on age, gender, current
and lifelong tobacco use, alcohol consumption, medical history, and
occupational history were collected by interview. Six hematological
measurements were evaluated: total WBC count, ALC, hematocrit, RBC
count, platelet count, and MCV. Total WBC counts and ALC were performed
using a Coulter T540 blood counter. Abnormal counts were confirmed.
Benzene metabolites in urine were measured by an isotope dilution gas
chromatography/mass spectometry assay. Correlation analyses were
performed with Spearman rank order correlation. The Wilcoxon rank sum
test was used to test for hematological differences.

Mean (standard deviation) years of occupational exposure to benzene were
6.3 (4.4) with a range of 0.7-16 years. The median 8-hour TWA benzene
exposure concentration for all exposed workers was 31 ppm (99 mg/m3).
Exposure to toluene and xylene was <= 0.2 ppm (0.6 mg/m3) in all groups.
The exposed group was subdivided into two equal groups of 22-one group
comprising workers who were exposed to greater than the median
concentration and the other containing those exposed to less than the
median concentration. The median (range) 8-hour TWA exposure
concentration was 13.6 (1.6-30.6) ppm (43.4 [5.1-97.8] mg/m3] for the
low-exposure group and 91.9 (31.5-328.5) ppm (294 [101-1049] mg/m3) for
the high-exposure group. A subgroup of the low-exposure group composed
of 11 individuals who were not exposed to >31 ppm (100 mg/m3) at any
time during the monitoring period was also examined in some comparisons.
The median (range) 8-hour TWA exposure of these individuals was 7.6
(1-20) ppm (24 [3.2-64] mg/m3). The urinary concentrations of the
metabolites phenol, muconic acid, hydroquinone, and catechol were all
significantly correlated with measured benzene exposure.

All six blood parameters measured were significantly different in the
high-benzene exposure group as compared to controls. ALC, WBC count, RBC
count, hematocrit, and platelets were all significantly decreased, and
MCV was significantly increased. The ALC was reduced from 1.9 x 103/µL
blood in controls to 1.6 x 103/µL (p<0.01) in the <31 ppm (99 mg/m3)
group and to 1.3 x 103/µL (p<0.001) in the group exposed to >31 ppm
benzene. In the subgroup of 11 workers exposed to a median 8-hour TWA of
7.6 ppm (24 mg/m3) benzene, the ALC (1.6 x 103/µL) was also
significantly reduced (p=0.03). The RBC and platelet counts were also
significantly reduced in the <31 ppm exposure group, but only ALC was
significantly different in the low-exposure subgroup. The fact that no
other measured blood cell parameters were significantly different in
this subgroup suggests that ALC was the most sensitive measure of
benzene hematotoxicity and that this exposure level (median 8-hour TWA
of 7.6 ppm) may be at the low end of the range of benzene exposures
eliciting hematotoxic effects in humans. 

ALC is also thought to have a potential role as a "sentinel" effect for
a cascade of early hematological and related biological changes that
might be expected to result in the more profound examples of benzene
poisoning observed in other cohorts of the National Cancer
Institute/Chinese Academy of Preventive Medicine study, as described by
Dosemeci et al. (1996). That ALC depletion is accompanied by
gene-duplicating mutations in somatic cells under the same range of
exposure conditions suggests that benzene can cause repeated damage to
longer-lived stem cells in human bone marrow, further implicating the
compound as etiologically important in the onset of benzene-associated
leukemia. This finding underlines the importance of basing public health
concern for benzene on a toxicological effect that is representative of
the earliest biological changes induced by the compound.

BMD modeling of the ALC exposure-response data from Rothman et al.
(1996) was done using U.S. EPA's Benchmark Dose Modeling Software
(version 1.20). The data are rather supralinear, that is, the change in
ALC per unit change in exposure decreases with increasing exposure;
therefore, in order to fit the data with one of the available continuous
models, the exposure levels were first transformed according to the
equation d’ = ln(d+1). Then the exposure-response data were fitted
using the continuous linear model, which provided a good fit (p=0.54). A
two-degree polynomial and a power model also fit the data, but the
linear model was selected because it is the most parsimonious. The
parameters were estimated using the method of maximum likelihood. A
constant variance model was used.

In the absence of a clear definition for an adverse effect for this
continuous endpoint, a default benchmark response of one standard
deviation change from the control mean was selected, as suggested in
EPA's draft Benchmark Dose Technical Guidance Document (U.S. EPA, 2000).
This default definition of a benchmark response for continuous endpoints
corresponds to an excess risk of approximately 10% for the proportion of
individuals below the 2nd percentile (or above the 98th percentile) of
the control distribution for normally distributed effects (see U.S. EPA,
2000). A 95% lower confidence limit (BMCL) on the resulting BMC was
calculated using the likelihood profile method. Transforming the results
back to the original exposure scale yields a BMC of 13.7 ppm (8-hr TWA)
and a BMCL of 7.2 ppm (8-hr TWA).

As suggested in the draft technical guidance document (U.S. EPA, 2000),
the BMCL is chosen as the point of departure for the RfC derivation. An
adjusted BMCL is calculated by converting ppm to mg/m3 and adjusting the
8-hour TWA occupational exposure to an equivalent continuous
environmental exposure. The BMCL is first converted to mg/m3 using the
molecular weight of 78.11 for benzene and assuming 25ºC and 760 mm Hg:
7.2 ppm x 78.11/24.45 = 23.0 mg/m3. The converted value is then adjusted
from the 8-hour occupational TWA to a continuous exposure concentration
using the default respiration rates (U.S. EPA, 1994): BMCLADJ = 23.0
mg/m3 x (10 m3/20 m3) x 5 days/7 days = 8.2 mg/m3.

The RfC is then derived by dividing the adjusted BMCL by the overall UF
of 300: RfC = BMCLADJ/UF = 8.2 mg/m3 ÷ 300 = 3 x 10-2 mg/m3. The
overall UF of 300 comprises a UF of 3 for effect-level extrapolation, 10
for intraspecies differences (human variability), 3 for
subchronic-to-chronic extrapolation, and 3 for database deficiencies
(see Section I.B.3).

For comparison, an RfC was also calculated based on the LOAEL of 7.6 ppm
(8-hr TWA) from the Rothman et al. (1996) study. Converting the units
and adjusting for continuous exposure as above results in a LOAELADJ of
8.7 mg/m3. The LOAELADJ is then divided by an overall UF of 1000 to
obtain the RfC: 8.7 mg/m3 ÷ 1000 = 9 x 10-3 mg/m3. The combined UF of
1000 represents UFs of 10 to account for the use of a LOAEL because of
the lack of an appropriate NOAEL, 10 for intraspecies differences in
response (human variability), 3 for subchronic-to-chronic extrapolation,
and 3 for database deficiencies. The value of 9 x 10-3 mg/m3 is in good
agreement with the RfC of 3 x 10-2 mg/m3 calculated from the BMC. 

A comparison RfC derivation based on BMD modeling of hematological data
from the Ward et al. (1985) subchronic experimental animal inhalation
study was also conducted. The Ward study was selected because it used a
relatively long inhalation exposure duration and an adequate number of
animals, and it provided dose-response data. Ward et al. exposed male
and female CD-1 mice and Sprague-Dawley rats to 0, 1, 10, 30 or 300 ppm
(0, 3.2, 32, 96 or 960 mg/m3) benzene, 6 hours/day, 5 days/week for 91
days and measured various hematological endpoints. The study identified
both a LOAEL of 300 ppm and a NOAEL of 30 ppm. The male mouse appeared
to be the most sensitive sex/species in this study. The
exposure-response relationships for the different hematological
endpoints for the male mouse were modeled using a BMD modeling approach
and decreased hematocrit (i.e., volume percentage of erythrocytes in
whole blood) was chosen as the critical effect.

U.S. EPA's Benchmark Dose Modeling Software (version 1.20) was used for
the modeling. An assumption of constant variance was used, although the
test for homogeneity of the variances failed. The continuous linear,
polynomial, and power models all resulted in the same BMC and BMCL
estimates; however, the linear model had better results for the fit
statistics. The linear model had a p-value of 0.09, which is of
borderline adequacy (the draft technical guidance document [U.S. EPA,
2000] recommends a p-value of >= 0.1), and the other models had p-values
of 0.04. Thus the continuous linear model was selected. The parameters
were estimated using the method of maximum likelihood.

In the absence of a clear definition for an adverse effect for this
continuous endpoint, a default benchmark response of one standard
deviation from the control mean was selected, as suggested in the draft
technical guidance document (U.S. EPA, 2000). The software uses the
estimated standard deviation. A 95% lower confidence limit (BMCL) on the
resulting BMC was calculated using the likelihood profile method. A BMC
of 100.7 ppm and a BMCL of 85.0 ppm 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
hematological 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. Therefore, there
is some uncertainty about the actual shape of the exposure-response
curve in the region of the benchmark response 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 to the human ALC results. Total WBC counts were reported and
exhibited the largest percent change in response between the control and
the 300 ppm group; however, the data for this endpoint also had
substantial variance, and because the benchmark response used for this
analysis is a function of the standard deviation, WBC count did not
yield the lowest BMC estimate. The actual lowest BMC estimates were
obtained for increased mean cell hemoglobin (MCH) (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 hematocrit. The results are also similar for
total blood hemoglobin (BMC = 104 ppm, BMCL = 88 ppm). RBC count results
were in between those for MCV and MCH and those for hematocrit and total
hemoglobin; however, the model fits were not adequate for the RBC data
and, thus, the RBC results have more uncertainty.

To derive the RfC, the BMCL is used as the point of departure, as
suggested in the draft Benchmark Dose Technical Guidance Document (U.S.
EPA, 2000). For conversion of the inhalation exposures across species,
ppm equivalence was assumed; this is identical to using EPA's inhalation
dosimetry methodology with Regional Gas Dose Ratio for the respiratory
tract region (RGDRr) = 1 (U.S. EPA, 1994). The BMCL is first converted
to mg/m3 using the molecular weight of 78.11 for benzene and assuming
25ºC and 760 mm Hg: BMCL (mg/m3) = 85.0 ppm x 78.11/24.45 = 272 mg/m3.
The converted value is then adjusted to an equivalent continuous
exposure: BMCLADJ = 272 mg/m3 x (6 hrs/24 hrs) x 5 days/7 days = 48.5
mg/m3.

The RfC is then obtained by dividing the adjusted BMCL by the overall UF
of 1000: RfC = 48.5 mg/m3 ÷ 1000 = 5 x 10-2 mg/m3. The overall UF of
1000 comprises a UF of 3 for effect-level extrapolation, 3 for
interspecies extrapolation (inhalation), 10 for intraspecies
differences, 3 for subchronic-to-chronic extrapolation, and 3 for
database deficiencies (see Section I.B.3). This value is in good
agreement with the RfC of 3 x 10-2 mg/m3 calculated from the BMC from
the Rothman et al. (1996) human study.

For further comparison, an RfC was also calculated, based on the NOAEL
of 30 ppm from the Ward et al. (1985) study. Converting the units and
adjusting for continuous exposure as above results in a NOAELADJ of 17.1
mg/m3. The NOAELADJ is then divided by an overall UF of 300 to obtain
the RfC: 17.1 mg/m3 ÷ 300 = 6 x 10-2 mg/m3. The combined UF of 300
represents a UF of 3 for interspecies extrapolation (inhalation), 10 for
intraspecies differences, 3 for subchronic-to-chronic extrapolation, and
3 for database deficiencies. The value of 6 x 10-2 mg/m3 is also in good
agreement with the RfC of 3 x 10-2 mg/m3 calculated from the BMC from
the Rothman et al. (1996) human study.

It should be noted, however, that other experimental animal studies have
reported significant hematological effects at benzene exposures of 10-25
ppm, which are lower than the NOAEL of 30 ppm from the Ward et al.
(1985) study. These studies have insufficient data for dose-response
modeling, and they used shorter exposure durations and/or fewer
experimental animals than did the Ward et al. (1985) study; nonetheless,
they observed statistically significant hematological effects at 10–25
ppm. Baarson et al. (1984), for example, exposed male C57BL/6J mice
(five/group) to 10 ppm benzene, 6 hours/day, 5 days/week, for 178 days
and observed statistically significant reductions in blood lymphocytes
at each of the three monitoring time points (32, 66, and 178 days) when
compared to controls. The magnitude of the reduction in lymphocytes
ranged from about 53% at 32 days to about 68% at 178 days. Cronkite et
al. (1985) exposed male and female C57BL/6 BNL mice to various
concentrations of benzene 6 hours/day, 5 days/week for 2 weeks and
observed no decrease in blood lymphocytes at 10 ppm, but they did
observe a statistically significant reduction of about 21% at 25 ppm as
compared to controls (5–10 mice/group). Thus, lower RfCs than those
calculated above for the Ward et al. (1985) study are possible, based on
other experimental animal results. In the most extreme case, using a
LOAEL of 10 ppm and an overall UF of 3000 yields a LOAELADJ of 5.7 mg/m3
and an RfC of 2 x 10-3 mg/m3.

__I.B.3. Uncertainty and Modifying Factors (Inhalation RfC)

UF = 300 for the BMCL from the Rothman et al. (1996) study.

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. U.S. EPA is planning to develop guidance for applying an effect
level extrapolation factor to a BMD. In the interim, a factor of 3 will
be used in this analysis (see Section I.A.3). For a more serious effect,
a larger factor, such as 10, might be selected. Second, a factor of 10
was used for intraspecies differences in response (human variability) as
a means of protecting potentially sensitive human subpopulations. Third,
a UF of 3 for subchronic-to-chronic extrapolation was applied (see
Section I.A.3). Finally, a UF of 3 was chosen to account for database
deficiencies, because no two-generation reproductive and developmental
toxicity studies for benzene are available. Therefore, an overall UF of
3 x 10 x 3 x 3 = 300 is used to calculate the RfC.

For the comparison analysis based on the Rothman et al. (1996) LOAEL,
the following UFs were selected: a factor of 10 for use of a LOAEL due
to lack of an appropriate NOAEL, a factor of 10 for intraspecies
variability, a factor of 3 for subchronic-to-chronic extrapolation, and
a factor of 3 for database deficiencies. Hence, an overall UF of 10 x 10
x 3 x 3 = 1000 was used in the comparison analysis.

For the comparison analysis based on the BMCL calculated from BMD
modeling of the male mouse data from the Ward et al. (1985) subchronic
inhalation study, the following UFs were used: a UF of 3 for
effect-level extrapolation, which is analogous to the LOAEL-to-NOAEL
extrapolation factor, because the BMC is considered an adverse effect
level; a UF of 3 for interspecies extrapolation for inhalation studies;
a UF of 10 for intraspecies variability; and a UF of 3 for database
deficiencies. In addition, a partial UF of 3 was used to extrapolate
from subchronic to chronic exposure. This partial value was selected
based on the observation that hematological fluctuations such as
reductions in RBCs and WBCs in the high-dose mice were noted at interim
sacrifice (14 days) as well as at termination (91 days), suggesting that
the responses occurred early in the exposure cycle and then remained
comparatively unchanged. Thus, an overall UF of 3 x 3 x 10 x 3 x 3 =
1000 was used in this comparison analysis.

Finally, for the comparison analysis based on the NOAEL from the Ward et
al. (1985) subchronic inhalation study, the following UFs were used: 3
for interspecies extrapolation for inhalation studies, 10 for
intraspecies variability, 3 for database deficiencies, and 3 for
subchronic-to-chronic extrapolation, as above. Therefore, an overall UF
of 300 was used in this comparison analysis.

MF = None. No modifying factor was considered necessary.

__I.B.4. Additional Studies/Comments (Inhalation RfC)

Benzene is toxic by all routes of administration. Hematotoxicity and
immunotoxicity have been consistently reported to be the most sensitive
indicators of noncancer toxicity in both humans and experimental
animals, and these effects have been the subject of several reviews
(Aksoy, 1989; Goldstein, 1988, Snyder et al., 1993; Ross, 1996; U.S.
EPA, 2002). The bone marrow is the target organ for the expression of
benzene hematotoxicity and immunotoxicity. Neither gastrointestinal
effects from oral exposure nor pulmonary effects due to inhalation
exposure have been reported.

Chronic exposure to benzene results in progressive deterioration in
hematopoietic function. Anemia, leukopenia, lymphocytopenia,
thrombocytopenia, pancytopenia, and aplastic anemia have been reported
after chronic benzene exposure (Aksoy, 1989; Goldstein, 1988). In an
earlier follow-up study of benzene-exposed workers, Aksoy et al. (1972)
reported that 8 of 32 workers who had been diagnosed with pancytopenia
died, mainly from infection and bleeding. In contrast to these blood
cellularity depression effects, benzene is also known to induce bone
marrow hyperplasia. Acute myelogenous leukemia has been frequently
observed in studies of human cohorts exposed to benzene, and there is
evidence linking benzene exposure to several other forms of leukemia.
Whether the hematotoxic/immunotoxic effects of benzene exposure and its
carcinogenic effects are due to a common mechanism is not yet known.
This is in part due to the fact that although the bone marrow depressive
effects of exposure to benzene in humans can be readily duplicated in
several experimental animal model systems, a suitable experimental
animal system for the induction of leukemia has not been found. The
hematotoxicity/immunotoxicity effects of benzene exposure lead to
significant health effects apart from potential induction of leukemia,
as several deaths due to aplastic anemia have been reported (ATSDR,
1997). 

Leukocytopenia has been consistently shown to be a more sensitive
indicator of benzene toxicity in experimental animal systems than
anemia, and lymphocytopenia has been shown to be an even more sensitive
indicator of benzene toxicity than overall leukocytopenia (Snyder et
al., 1980, Ward et al., 1985; Baarson et al., 1984). Rothman et al.
(1996) also found that a decrease in ALC was the most sensitive
indicator of benzene exposure in a group of workers. Ward et al. (1996)
observed a strong relationship between benzene exposure and decreased
WBC counts in a rubber worker cohort, but no significant relationship
with RBC counts was found.

Bogardi-Sare et al. (2000) found that exposure to benzene concentrations
of less than 15 ppm can induce depression of circulating B-lymphocytes.
Dosemeci et al. (1996) were able to demonstrate the presence of benzene
poisoning (WBC <4000 cells/mm3 and platelet count <80,000/mm3) at levels
of exposure in the 5–19 ppm range. 

As is the case with many other organic solvents, benzene has been shown
to produce neurotoxic effects in test animals and humans after
short-term exposures to relatively high concentrations (U.S. EPA, 2002).
The neurotoxicity of benzene, however, has not been extensively studied,
and no systematic studies of the neurotoxic effects of long-term
exposure have been conducted. Additionally, there is some evidence from
human epidemiologic studies of reproductive and developmental toxicity
of benzene, but the data did not provide conclusive evidence of a link
between exposure and effects (U.S. EPA, 2002). Some test animal studies
provide limited evidence that exposure to benzene affects reproductive
organs; however, these effects were limited to high exposure
concentrations that exceeded the maximum tolerated dose (U.S. EPA,
2002). Results of inhalation studies conducted in test animals are
fairly consistent across species and have demonstrated that at
concentrations of greater than 150 mg/m3 (47 ppm) benzene is fetotoxic
and causes decreased fetal weight and/or minor skeletal variants (U.S.
EPA, 2002). Exposure of mice to benzene in utero has also been shown to
cause changes in the hematogenic progenitor cells in fetuses, 2-day
neonates, and 6 week-old adults (Keller and Snyder, 1986, 1988). 

For more detail on Susceptible Populations, exit to   HYPERLINK
"http://www.epa.gov/iris/toxreviews/0276-tr.pdf" \l "page=124"  the
toxicological review, Section 4.4  (PDF). 

__I.B.5. Confidence in the Inhalation RfC

Study — Medium

Database — Medium

RfC — Medium

The overall confidence in this RfC assessment is medium. The principal
study of Rothman et al. (1996) was well conducted, and the availability
of good-quality human data for a sensitive endpoint eliminates the
uncertainty associated with basing the RfC on experimental animal data.
In addition, the RfC of 3 x 10-2 mg/m3 obtained from the BMD modeling
results from the Rothman et al. (1996) study is in good agreement with
the value of 9 x 10-3 mg/m3 based on the LOAEL. The RfC is also in good
agreement with the values of 5 x 10-2 mg/m3 and 6 x 10-2 mg/m3 based on
the BMC and the NOAEL, respectively, from the Ward et al. (1985)
subchronic rodent inhalation study. This consistency in results provides
increased confidence in the RfC. 

With continuous endpoints such as hematological parameters, there is
uncertainty about when a change in a parameter that has inherent
variability becomes an adverse effect. Other uncertainties explicitly
recognized in the quantitative derivation include intraspecies
variability (to accommodate sensitive human subgroups),
subchronic-to-chronic extrapolation, and database deficiencies due to
the lack of two-generation reproductive and well-conducted developmental
toxicity studies for benzene. 

For more detail on Characterization of Hazard and Dose Response, exit to
  HYPERLINK "http://www.epa.gov/iris/toxreviews/0276-tr.pdf" \l
"page=156"  the toxicological review, Section 6  (PDF). 

__I.B.6. EPA Documentation and Review of the Inhalation RfC

Source Document — U.S. EPA, 2002. 

This assessment was peer reviewed by external scientists as well as in
response to public comments. Their comments have been evaluated
carefully and incorporated in the finalization of this IRIS Summary. The
  HYPERLINK "http://www.epa.gov/iris/supdocs/benzene_nc-pr.pdf"  peer
review document  (12 pages, 135 Kbytes) is available in Adobe PDF
format.

Other EPA Documentation — None

Date of Agency Consensus — January 23, 2002 

__I.B.7. EPA Contacts (Inhalation RfC)

Please contact the IRIS Hotline for all questions concerning this
assessment or IRIS, in general, at (202)566-1676 (phone), (202)566-1749
(FAX) or   HYPERLINK "mailto:hotline.iris@epa.gov"  hotline.iris@epa.gov
 (internet address). 

  HYPERLINK "http://www.epa.gov/iris/subst/0276.htm" \l
"content#content"  Top of page 

_II.  Carcinogenicity Assessment for Lifetime Exposure

Substance Name — Benzene

CASRN — 71-43-2

Last Revised — 01/19/2000

Section II provides information on three aspects of the carcinogenic
assessment for the substance in question; the weight-of-evidence
judgment of the likelihood that the substance is a human carcinogen, and
quantitative estimates of risk from oral exposure and from inhalation
exposure. The quantitative risk estimates are presented in three ways.
The slope factor is the result of application of a low-dose
extrapolation procedure and is presented as the risk per (mg/kg)/day.
The unit risk is the quantitative estimate in terms of either risk per
ug/L drinking water or risk per ug/cu.m air breathed. The third form in
which risk is presented is a drinking water or air concentration
providing cancer risks of 1 in 10,000, 1 in 100,000 or 1 in 1,000,000.
The rationale and methods used to develop the carcinogenicity
information in IRIS are described in The Risk Assessment Guidelines of
1986 (EPA/600/8-87/045) and in the IRIS Background Document. IRIS
summaries developed since the publication of EPA's more recent Proposed
Guidelines for Carcinogen Risk Assessment also utilize those Guidelines
where indicated (Federal Register 61(79):17960-18011, April 23, 1996).
Users are referred to Section I of this IRIS file for information on
long-term toxic effects other than carcinogenicity. 

_II.A. Evidence for Human Carcinogenicity

__II.A.1. Weight-of-Evidence Characterization

Benzene is classified as a "known" human carcinogen (Category A) under
the Risk Assessment Guidelines of 1986. Under the proposed revised
Carcinogen Risk Assessment Guidelines (U.S. EPA, 1996), benzene is
characterized as a known human carcinogen for all routes of exposure
based upon convincing human evidence as well as supporting evidence from
animal studies. (U.S. EPA, 1979, 1985, 1998; ATSDR, 1997). 

Epidemiologic studies and case studies provide clear evidence of a
causal association between exposure to benzene and acute nonlymphocytic
leukemia (ANLL) and also suggest evidence for chronic nonlymphocytic
leukemia (CNLL) and chronic lymphocytic leukemia (CLL). Other neoplastic
conditions that are associated with an increased risk in humans are
hematologic neoplasms, blood disorders such as preleukemia and aplastic
anemia, Hodgkin's lymphoma, and myelodysplastic syndrome (MDS). These
human data are supported by animal studies. The experimental animal data
add to the argument that exposure to benzene increases the risk of
cancer in multiple species at multiple organ sites (hematopoietic, oral
and nasal, liver, forestomach, preputial gland, lung, ovary, and mammary
gland). It is likely that these responses are due to interactions of the
metabolites of benzene with DNA (Ross, 1996; Latriano et al., 1986).
Recent evidence supports the viewpoint that there are likely multiple
mechanistic pathways leading to cancer and, in particular, to
leukemogenesis from exposure to benzene (Smith, 1996). 

__II.A.2. Human Carcinogenicity Data

Benzene is a known human carcinogen based upon evidence presented in
numerous occupational epidemiological studies. Significantly increased
risks of leukemia, chiefly acute myelogenous leukemia (AML), have been
reported in benzene-exposed workers in the chemical industry,
shoemaking, and oil refineries. 

The following epidemiologic studies briefly described are the key
studies that support the weight-of-evidence classification that exposure
to benzene is causally related to an increase in the risk of cancer,
specifically leukemia. 

Aksoy et al. (1974) reported effects of benzene exposure among 28,500
Turkish workers employed in the shoe industry. The mean duration of
employment was 9.7 years (range 1 to 15 years) and the mean age was 34.2
years. Peak exposure to benzene was reported to be 210 to 650 ppm.
Twenty-six cases of leukemia and a total of 34 leukemias or preleukemias
were observed, corresponding to an incidence of 13/100,000 (by
comparison to 6/100,000 for the general population). A follow-up
analysis of the study (Aksoy, 1980) reported eight additional cases of
leukemia as well as evidence suggestive of increases in other
malignancies. This case study lacks detailed information on personal
exposure to benzene and potential exposure to other chemicals, a
well-defined comparison population, and control of confounding
variables. 

Infante et al. (1977b), in a retrospective cohort mortality study,
examined the leukemogenic effects of benzene exposure in 748 white male
workers exposed at least 1 day while employed in the manufacture of
rubber products. Exposure occurred from 1940 to 1949 and vital status
was obtained through 1975. A statistically significant increased risk of
leukemia (7 observed, 1.48 expected; p < .002) was found by comparison
of observed leukemia deaths in this cohort with those expected based
upon general U.S. population death rates. The risk of leukemia was said
by the authors to be potentially understated since follow-up was only
75% complete. According to the authors, there was no evidence of solvent
exposure other than benzene. No effort was made to evaluate individual
exposures to benzene for the purpose of doing a dose-response analysis.
The main criticism of this study, as well as its later updates, is the
small size of the cohort. 

In an extension and elaboration of the analysis done by Infante et al.
(1977b), Rinsky et al. (1981) reported seven deaths from leukemia in
this same cohort after achieving a 98% vital status ascertainment
through June 1975. Forty additional deaths from all causes were
reported, but no new leukemia deaths. Again, the risk of death from
leukemia was statistically significant (standardized mortality ratio
[SMR] was 560 based upon 7 leukemia deaths, p < .001). Some 437 members
of the cohort were exposed for less than 1 year. Those who received 5 or
more years of exposure exhibited an SMR of 2100, based upon 5 leukemia
deaths versus 0.25 expected (p < .01). All seven leukemia cases were of
the myelogenous or monocytic cell type. Four additional deaths from
leukemia were also noted but could not be added to the total because
they did not fit the criteria for inclusion. The authors tried to
reconstruct past exposure to benzene at the two locations of this
company and found that in some areas of the plants airborne benzene
concentrations occasionally rose to several hundred parts per million,
but most often employee 8-hour time-weighted averages (TWA) fell within
the limits considered permissible at the time of exposure. No
dose-response analysis was attempted. 

In an updated version of the Rinsky et al. (1981) study, the same
authors examined a somewhat expanded cohort of 1165 nonsalaried white
men employed in the rubber hydrochloride department for at least 1 day
through December 1965 and followed to December 31, 1981 (Rinsky et al.,
1987). Followup was 98.6% complete. Again, a statistically significant
excess risk of leukemia was found for the total cohort (9 observed, 2.7
expected; p < 0.05). For the first time, individual measurements of
cumulative exposure in terms of ppm-years were generated for all members
of the cohort utilizing the historical air-sampling data discussed above
or interpolating estimates based on the existing data. SMRs for leukemia
ranged from a nonsignificant 109 (2 observed, 1.83 expected) at
cumulative exposures under 40 ppm-years to a statistically significant
SMR of 2339 (5 observed, 0.21 expected; p < .05) at 200 ppm-years or
more of exposure. The authors found significantly elevated risks of
leukemia at cumulative exposures less than the then equivalent current
standard for occupational exposure, which was 10 ppm over a 40-year
working lifetime. 

The Rinsky et al. (1981, 1987) study analyses, based upon the original
cohort of Pliofilm rubber workers studied by Infante et al. (1977b),
were selected by the Agency as the critical study for dose-response
analysis and for the quantitative estimation of cancer risk to humans.
The Rinsky et al. (1981, 1987) analyses show ample power, latency,
reasonably good estimates of exposure to benzene except prior to 1946,
few confounders, and a wide range of exposure to benzene from low levels
to high levels. Limitations include the small cohort size, reporting
only nine leukemia deaths with no estimates of risk according to cell
type. There remain questions about the estimation of personal exposure
to benzene, especially prior to 1946 when no measurements of airborne
benzene were made. And finally, at levels less than 200 ppm-years it is
not possible to determine leukemia risk in this cohort because of lack
of sensitivity of the data at low levels. 

Ott et al. (1978) observed a nonsignificantly increased risk of leukemia
(3 deaths) among 594 chemical workers exposed to benzene followed for at
least 23 years in a retrospective cohort mortality study. Benzene
exposures ranged from under 2 ppm to over 25 ppm 8-hour TWA. Bond et al.
(1986) updated this report by following this cohort an additional 9
years to the end of 1982 and adding an additional 362 exposed workers
not studied previously. The authors reported finding a nonsignificant
excess risk (SMR = 194) of deaths from leukemia based upon 4 cases. All
were diagnosed as myelogenous leukemias. The authors reported that this
represented a significant excess (4 observed versus 0.9 expected, p <
.011) for myelogenous leukemia based upon the International
Classification of Diseases and Causes of Death. It is not stated whether
these four deaths were acute or chronic. One additional death was
classified as a "pneumonia" death, but on the death certificate "acute
myelogenous leukemia" was noted as a significant contributing condition.
Cumulative exposure estimates ranged from 18 ppm-months to a high of
4211 ppm-months. The Bond et al. (1986) study has little power to detect
significant risk of leukemia at low doses. The authors also state that
their data should not be used for determining unit risk estimates
because of the small number of events, competing exposures to other
potentially hazardous materials, and the contribution of unquantified
brief exposures to benzene. 

Wong (1983, 1987) reported on the mortality of male chemical workers who
had been exposed to benzene for at least 6 months during the years 1946
to 1975. The study population of 4602 persons was drawn from seven
chemical plants and cumulative exposures to benzene were determined for
all subjects. The control subjects (3074 persons) held jobs at the same
plants for at least 6 months but were never subjected to benzene
exposure. Dose-dependent increases were seen in the risk of leukemia and
the risk of lymphatic and hematopoietic cancer. Chemical workers with a
cumulative exposure to benzene of 720 ppm-months were subject to a
borderline significant relative risk of 3.93 (p = .05) for lymphatic and
hematopoietic cancer. None of the leukemia deaths were of the acute
myeloid cell type, the type that was known to be associated with
exposure to benzene in other studies. The author further observed that
cumulative exposure, not peak exposure, was the major variable in
quantifying mortality risk from lymphopoietic cancer. The
Mantel-Haenszel chi-square for upward trend in risk of leukemia with
increasing cumulative exposure was significantly elevated at the 99%
level of confidence. Some of the limitations of this study include
imprecise historical industrial hygiene data, unusual distribution of
leukemia cell types, i.e., there were no acute cases of myelogenous
leukemia out of seven leukemia cases, and possible exposure of
comparison subjects to potentially carcinogenic solvents other than
benzene. 

The National Cancer Institute of the U.S. National Institutes of Health
and the Chinese Academy of Preventative Medicine have been conducting a
comprehensive epidemiological study of 74,828 benzene-exposed workers
employed from 1972 to 1987 in 672 factories in 12 cities of China
(Dosemeci et al., 1994; Hayes et al., 1996, 1997; Yin et al., 1987,
1989, 1994, 1996). A comparison group of workers consisting of 35,805
employees was assembled from non-benzene-exposed units of 69 of the same
factories and 40 factories elsewhere. Workers in a variety of jobs in
painting, printing, footwear, rubber, chemical, and other industries
were followed for vital status for an average period of time of less
than 12 years. Less than 0.3% were lost to follow-up. Employee work
histories were linked to benzene exposure data in order to derive
individual time-specific estimates for each worker (Dosemeci et al.,
1994). This large cohort mortality study produced a significantly
elevated risk of hematologic neoplasms (RR = 2.2, 95% C.I. = 1.1-4.2) in
workers exposed to benzene at an average level of less than 10 ppm. A
combination of ANLL and MDS produced a risk of 3.2 (95% C.I .=
1.0-10.1). For exposure to a sustained concentration of 25 ppm benzene,
the risk of ANLL and MDS increased to 7.1 (95% C.I. = 2.1-23.7). The
risk of other leukemias (other than ANLL), including chronic myeloid and
monocytic leukemia, was not significantly elevated (RR = 2.0).
Additionally, the risk of non-Hodgkin's lymphoma was significantly
elevated (RR = 4.2 with 95% C.I. = 1.1-15.9) for those with a sustained
exposure to benzene that occurred at least 10 years prior to diagnosis.
The authors concluded that benzene exposure "is associated with a
spectrum of hematologic neoplasms and related disorders in humans and
that risks for these conditions are elevated at average benzene-exposure
levels of less than 10 ppm." Limitations of this study include possible
concurrent exposures to many different chemicals found in the factories
where the benzene exposure occurred. There is a lack of reliable
exposure information in the early days of the observation period, when
only 3% of the exposure estimates were based on actual measurements. 

All of the epidemiological studies referred to above have some
methodological problems, i.e., confounding exposures, lack of sufficient
power, and other limitations, but the consistent excess risk of leukemia
across all of these studies argues that such problems could not be
entirely responsible for the elevated risks of cancer. Most of these
epidemiologic and case studies have been reviewed in peer-reviewed
publications (IARC, 1982; ATSDR, 1997; U.S. EPA, 1998). They provide
clear evidence of a causal association between exposure to benzene and
ANLL. The evidence is suggestive with respect to CNLL and CLL. 

The limitations of these studies, except for Rinsky et al. (1981, 1987),
preclude their use in quantitative risk estimation. This is further
discussed in the quantitative risk estimation sections (II.C.3 and
II.C.4). 

__II.A.3. Animal Carcinogenicity Data

Although human epidemiological studies provide the bulk of the evidence
reaffirming the classification of benzene as a category A, "known" human
carcinogen (U.S. EPA, 1979, 1985, 1998), many experimental animal
studies, both inhalation and oral, also support the evidence that
exposure to benzene increases the risk of cancer in multiple organ
systems including the hematopoietic system, oral and nasal cavities,
liver, forestomach, preputial gland, lung, ovary, and mammary gland. The
key animal studies that support the finding of an excess risk of
leukemia in humans from exposure to benzene by the inhalation route are
Maltoni et al. (1982, 1983, 1985, 1989), Cronkite et al. (1984, 1985,
1989), Snyder et al. (1988), and Farris et al. (1993); and by the oral
route, Huff et al. (1989), NTP (1986), and Maltoni et al. (1983, 1985,
1989). The details of these studies have been reviewed (ATSDR, 1997).
Studies on the carcinogenicity of benzene in rodents include inhalation
exposures to Sprague-Dawley rats, C57BL/6 mice, AKR mice, CD-1 mice, and
CBA mice; and gavage treatment of Sprague-Dawley rats, Wistar rats, F344
rats, RF/J mice, Swiss mice, and B6C3F1 mice (Cronkite et al., 1989;
Goldstein et al., 1982; Huff et al., 1989; Maltoni et al., 1983, 1988;
NTP, 1986; Snyder et al., 1980, 1982, 1984; Farris et al., 1993).
Inhalation concentrations ranged from 0 to 1000 ppm and gavage doses
ranged from 0 to 200 mg/kg. 

It is noted that in humans the cancer induced by benzene exposure is
predominantly acute nonlymphocytic leukemia, while in rodents
lymphocytic leukemia was observed in two series of experiments in
C57BL/6 mice (Snyder et al., 1980) and CBA/Ca mice (Cronkite et al.,
1989). The difference in induction of hematopoietic cancers in mice and
humans is not fully understood, but it may be related to
species-specific differences in hematopoiesis. Lymphocytes make up a
larger portion of the nucleated cells in mouse bone marrow than in human
bone marrow (Parmley, 1988) and could simply represent a larger target
cell population for benzene metabolites. The bone marrow, Zymbal gland,
and Harderian gland all contain peroxidases, which can activate phenols
to toxic quinones and free radicals. Sulfatases, which remove conjugated
sulfate and thus reform free phenols, are also present at high levels in
these target organs. The selective distribution of these two types of
enzymes in the body may explain the accumulation of free phenol,
hydroquinone, and catechol in the bone marrow and the resulting
differences in target organ toxicity of benzene metabolites in humans
and animals. The animal bioassay results may have some relevance to
human leukemia, but it should be emphasized that there is no
well-demonstrated and reproducible animal model for leukemia resulting
from benzene exposure. The mechanism of leukemia development following
exposure to benzene is not well understood (Low et al., 1989, 1995). 

__II.A.4. Supporting Data for Carcinogenicity 

The supporting evidence for the carcinogenic effects of exposure to
benzene comes from our current understanding of the metabolism and mode
of action (Stephens et al., 1994; Medinsky et al., 1996; Lee et al.,
1996; Valentine et al., 1996; Rothman, 1997). This is briefly summarized
below and reviewed in U.S. EPA (1998). 

It is generally agreed that the toxicity of inhaled benzene results from
its biotransformation to reactive species. Benzene is metabolized in the
liver by cytochrome P4502E1 (CYP2E1) to its major metabolites: phenol,
hydroquinone, and catechol. The intermediate benzene oxide can also
undergo ring opening to trans-trans muconic acid. Although there is a
scientific consensus that metabolism of benzene is required for
resultant toxicity and carcinogenic response, the role of a metabolite
or metabolites of benzene in producing these adverse effects is
controversial and more research data are needed to better define
sequelae of pathogenesis following exposure to benzene and its
metabolites. Current evidence indicates that benzene-induced
myelotoxicity and genotoxicity result from a synergistic combination of
phenol with hydroquinone, muconaldehyde, or catechol. 

Molecular targets for the action of these metabolites, whether acting
alone or in concert, include tubulin, histone proteins, topoisomerase
II, and other DNA-associated proteins. Damage to these proteins would
potentially cause DNA strand breakage, mitotic recombination,
chromosomal translocations, and malsegregation of chromosomes to produce
aneuploidy. If these effects took place in stem or early progenitor
cells, a leukemic clone with selective advantage to grow could arise as
a result of protooncogene activation, gene fusion, and suppressor-gene
inactivation. Epigenetic effects of benzene metabolites on the bone
marrow stroma, and perhaps the stem cells themselves, could then foster
development and survival of a leukemic clone. Since these plausible
events have not been conclusively demonstrated, this remains a
hypothesis (Smith, 1996). 

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_II.B. Quantitative Estimate of Carcinogenic Risk from Oral Exposure

__II.B.1. Summary of Risk Estimates

___II.B.1.1. Oral Slope Factor — 1.5 x 10-2 to 5.5 x 10-2 per
(mg/kg)/day

___II.B.1.2. Drinking Water Unit Risk — 4.4 x 10-7 to 1.6 x 10-6 per
(ug/L)

___II.B.1.3. Extrapolation Method — Linear extrapolation of human
occupational data

Drinking Water Concentrations at Specified Risk Levels: 

Risk Level 	Concentration 

E-4 (1 in 104)	102 µg/L to 103 µg/L

E-5 (1 in 105)	101 µg/L to 102 µg/L

E-6 (1 in 106)	100 µg/L to 101 µg/L

__II.B.2. Dose-Response Data (Carcinogenicity, Oral Exposure)

Tumor Type - leukemia

Test Species - human

Route - inhalation, occupational exposure

Reference - Rinsky et al., 1981, 1987; Paustenbach et al., 1993; Crump
1994; U.S. EPA, 1998; U.S. EPA, 1999. 

The quantitative oral unit risk estimate is an extrapolation from the
known inhalation dose-response to the potential oral route of exposure
documented in Section II.C. The inhalation risk estimate is reported as
a range, from 2.2 x 10-6 to 7.8 x 10-6 per µg/m3. No relevant data
exist in the published literature for oral absorption of benzene in
humans. Inhalation absorption is assumed to be about 50% while that of
oral is selected as 100% based upon a review of the relevant human and
animal literature (U.S. EPA, 1999). Absorption of benzene via the dermal
route of exposure is usually less than 1% of the applied dose and
therefore it is not considered to contribute significantly the oral risk
estimation. In the previous oral unit risk estimate it was assumed that
absorption was equal for both the inhalation and oral routes of exposure
(U.S. EPA, 1992). The inhalation unit risk range (per µg/m3) is first
converted to the oral slope factor, which is in units of risk per
µg/kg/day, by assuming a standard air intake of 20 m3/day, a standard
body weight of 70 kg for an adult human, and 50% absorption via
inhalation. The drinking water unit risk was then calculated from the
oral slope factor assuming a drinking water intake of 2 L/day. In
calculating the drinking water concentrations for specific risk levels,
the upper and lower end of the range round off to a single value. 

This assessment of the oral unit risk range replaces the previous oral
carcinogenicity assessment on IRIS dated April 1, 1992.

__II.B.3. Additional Comments (Carcinogenicity, Oral Exposure)

EPA's quantitative estimate for the cancer risk associated with
inhalation exposure to benzene was recently updated (U.S. EPA, 1998).
The new inhalation unit risk estimate is reported as a range, from 2.2 x
10-6 to 7.8 x 10-6 per µg/m3 (U.S. EPA, 1999b). To extrapolate to oral
risk, the inhalation unit risk range is first converted to units of dose
(µg/kg/day). Using the standard air intake factor of 20 m3/day, the
standard weight estimate of 70 kg, and the 50% absorption factor for
inhalation exposure given above, the dose from 1 µg/m3 continuous daily
exposure is: 

1 µg/m3*20 m3/day*0.5*(1/70) kg = 0.143 µg/kg/day 

The risk estimate range is then divided by this dose, to generate an
oral slope factor in units of inverse dose:

risk/(µg/kg/day) 	= 2.2 x 10-6/0.143 µg/kg/day to 7.8 x 10-6/0.143
µg/kg/day

= 1.54 x 10-5 to 5.45 x 10-5 per µg/kg/day

Assuming 100% oral absorption and a standard intake of 2 L/day, the
concentration in drinking water that would produce a dose of 1
µg/kg/day is: 

1 µg/kg/day*70 kg*(2 L/day)-1 = 35 µg/L 

Thus, the oral unit risk, in units of risk/(µg/L) would be: 

(1.54 x 10-5 to 5.45 x 10-5)/35 µg/L = 4.4 x 10-7 to 1.6 x 10-6/µg/L 

Note: This estimate is a risk factor for ingested benzene, and is not
sufficient to account for total exposure to drinking water. For
development of a drinking water safe concentration, the risk due to
inhalation of volatilized benzene from drinking water and to dermal
uptake must be added to the ingestion risk (Beavers et al., 1996;
Lindstrom et al., 1994). 

If one assumes a 20% respiratory absorption rate, the lowest value in a
group of subjects (range 20% to 50%) that was found in one study (Srbova
et al., 1950), then the oral unit risk range becomes 1.10 x 10-6 to 3.89
x 10-6. This may represent an upper bound on the risk range. The
standard values (20 m3/day, 70 kg, 2 L/day) used for the risk estimation
do not necessarily account for the population variability. 

The range of risk estimates of 4.4 x 10-7 to 1.6 x 10-6 /µg/L is
recommended, within which any value will have equal scientific
plausibility. The assumption is made that the leukemia effect is
dependent on the absorbed dose. For inhalation, the metabolized dose is
assumed to be 50% of the inhaled dose. This conclusion is supported by
studies in humans (Pekari et al., 1992; Hunter, 1966; Hunter, 1968;
Hunter and Blair, 1972; Nomiyama and Nomiyama, 1974; Srbova et al.,
1950; Teisinger et al., 1952, as cited in Fiserova-Bergerova et al.,
1974; Yu and Weisel, 1998) and by a pharmacokinetic model developed by
Bois et al. (1996). In the absence of data in humans regarding the
fraction of orally-ingested benzene that is metabolized, data from mice
and rats (Sabourin et al., 1987) suggests that there is a complete
absorption of the dose received by corn oil gavage and
intraperitoneally. 

__II.B.4. Discussion of Confidence (Carcinogenicity, Oral Exposure)

The most useful available human epidemiological data for evaluation of
the risk of cancer from exposure to benzene comes from occupational
inhalation exposure studies (Rinsky et al., 1981, 1987). There are few
human data regarding oral exposure to benzene. Route-to-route
extrapolation is justified because similar toxic effects are observed in
animals through either the oral or inhalation route of exposure to
benzene (ATSDR, 1997) and toxicokinetic data available from animal
studies (Gerrity et al., 1990). Experimental animal data also
demonstrate that benzene is metabolized to the same products whether it
is inhaled or ingested. Therefore, it is reasonable to extrapolate from
inhalation dose-response to estimate an equivalent oral dose-response. 

A rigorous method for route-to-route extrapolation that involves the
development of a pharmacokinetic model to predict the concentration of
the ultimate carcinogen in bone marrow has been proposed but has not
been validated (Smith and Fanning, 1997). Furthermore, the nature of the
distribution of benzene metabolites to the bone marrow is not well
understood. The chemical species responsible for the induction of
leukemia in animals and humans may involve more than one metabolite
(Smith, 1996). 

The absorption efficiencies across pulmonary and gastrointestinal
barriers provide an informed basis to adopt reasonable values for
benzene absorption. The oral slope factor is derived from the inhalation
slope factor currently documented in the IRIS database (Section II.C).
No relevant oral benzene exposure data on humans are available, but it
is known that complete gastrointestinal absorption occurs in the rat and
mouse study as reported by Sabourin et al (1987); it is reasonable to
assume complete absorption in humans. However, it is clear from numerous
studies of pulmonary absorption in humans that absorption of benzene via
the inhalation route is incomplete. There is a general consensus in the
literature supporting the use of a 50% absorption via inhalation and not
using default assumptions that assume both exposure routes have
equivalent absorption efficiencies. Based upon several inhalation
studies, EPA has judged an absorption factor of 50% to be the most
scientifically sound. 

In the absence of evidence to the contrary, key studies support the
reasonableness of extrapolating from inhalation to oral cancer risk. The
calculations use standard EPA conversion factors for air and water
intake and informed assumptions about the amount of absorption of
benzene from oral and inhalation exposure. 

A substantial literature provides information on pulmonary absorption in
humans. The animal study selected for this assessment provides excellent
information in two species for both inhalation and oral absorption.
However, data on oral absorption from drinking water exposure would be a
useful addition. 

While the human data demonstrate good agreement indicating that
approximately one- half of inhaled benzene is absorbed into the
bloodstream at exposure concentrations between 1 and 100 ppm,
considerable interindividual variability was observed in all studies
that reported on multiple subjects. Many factors, including activity
level, pulmonary health, and metabolic clearance, are likely to
influence the amount of benzene actually taken up in a diverse
population exposed by the inhalation route. To date, characterization of
the extent of variability is limited. 

The simple absorption ratio approach taken to route-to-route
extrapolation here cannot account for differences in disposition of
benzene after it crosses the pulmonary or gastrointestinal barrier.
First-pass metabolism of ingested benzene may have significant effects
on the dose of benzene metabolites that reaches the target bone marrow
cells (Sabourin et al., 1989). Leukemogenic metabolites may be produced
more efficiently after ingestion, but on the other hand, rapid clearance
of benzene and metabolites after ingestion may be a mitigating factor.
The data are inadequate to address these questions for humans at this
time, but a variety of biomarkers of benzene exposure can help to
address questions of internal dose of benzene metabolites. Biomarker
data, together with further development of PBPK models, using human data
to define parameters wherever possible, may provide improved dose
metrics for benzene risk assessment in the near future.

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_II.C. Quantitative Estimate of Carcinogenic Risk from Inhalation
Exposure

__II.C.1. Summary of Risk Estimates

___II.C.1.1. Air Unit Risk 

A range of 2.2 x 10-6 to 7.8 x 10-6 is the increase in the lifetime risk
of an individual who is exposed for a lifetime to 1 µg/m3 benzene in
air. 

___II.C.1.2. Extrapolation Method: 

Low-dose linearity utilizing maximum likelihood estimates (Crump, 1992,
1994). 

Air Concentrations at Specified Risk Levels: 

Risk Level	Concentration

E-4 (1 in 10,000)	13.0 to 45.0 µg/m3

E-5 (1 in 100,000)	1.3 to 4.5 µg/m3

E-6 (1 in 1,000,000)	0.13 to 0.45 µg/m3

__II.C.2. Dose-Response Data for Carcinogenicity, Inhalation Exposure

Tumor Type — Leukemia

Test Species -- Humans

Route — Inhalation

References -- Rinsky et al., 1981, 1987; Paustenbach et al., 1993; Crump
and Allen, 1984; Crump, 1992, 1994; U.S. EPA, 1998. 

__II.C.3. Additional Comments (Carcinogenicity, Inhalation Exposure)

The Pliofilm workers of Rinsky et al. (1981, 1987) provide the best
published set of data to date for evaluating human cancer risks from
exposure to benzene. Compared to the published studies of Ott et al.
(1978), Bond et al. (1986), and Wong (1987), this cohort has fewer
reported co-exposures to other potentially carcinogenic substances in
the workplace that might confound risk analysis for benzene. This cohort
also provides a greater range of exposures than those of Ott et al.
(1978), Bond et al. (1986), and Wong (1987). The Rinsky et al. data were
used for developing the unit cancer risk by Crump (1992, 1994).
Differences in the unit risk estimates, in addition to the choice of
model used, stem largely from differences in the exposure estimates and
the dose-response model used. 

Although the ongoing Chinese cohort studies (Dosemeci et al., 1994;
Hayes et al., 1996, 1997; Yin et al., 1987, 1989, 1994, 1996) provide a
large data set and perhaps may provide information in the future to
better characterize risk of cancer at low dose exposure, their use in
the derivation of risk estimates remains problematic at present for the
reasons cited in Section II.A.2. 

The two most important determinants of the magnitude of the unit risk
number are the choice of extrapolation model to be used to estimate risk
at environmental levels of exposure and the choice of the exposure
estimates to which the Pliofilm workers (Rinsky et al., 1981, 1987) were
subjected. Crump (1992, 1994) presented 96 unit risk calculation
analyses by considering different combinations of the following factors:
(1) different disease endpoints, (2) additive or multiplicative models,
(3) linear/nonlinear exposure-response relationships, (4) two different
sets of exposure measurements (Crump and Allen [1984] vs. exposure
estimates by Paustenbach et al. [1993]) and (5) cumulative or weighted
exposure measurements. The unit risk estimates range from 8.6 x 10-5 to
2.5 x 10-2 at 1 ppm (3200 µg/m3) of benzene air concentration (Crump,
1992, 1994). 

The risk estimates would fall into the lower range if a sublinear
exposure response model were found to be more plausible. However, the
shape of the exposure dose-response curve cannot be considered without a
better understanding of the biological mechanism(s) of benzene-induced
leukemia. Understanding of the mechanisms by which exposure to benzene
and its metabolites exert their toxic and carcinogenic effects remains
uncertain (U.S. EPA, 1998). It is likely that more than one mechanistic
pathway may be responsible for the toxicity of benzene contributing to
the leukemogenic process. Not enough is known to determine the shape of
the dose-response curve at environmental levels of exposure and to
provide a sound scientific basis to choose any particular extrapolation
model to estimate human cancer risk at low doses. In fact, recent data
(Hayes et al., 1997) suggest that because genetic abnormalities appear
at low exposures in humans, and the formation of toxic metabolites
plateaus above 25 ppm, the dose-response curve could be supralinear
below 25 ppm. Given this, EPA believes that use of a linear
extrapolation model as a default approach is appropriate. 

When a linear model was employed, the choice of cancer unit risk
estimates narrows to a range between 7.1 x 10-3 and 2.5 x 10-2 at 1 ppm
(2.2 x 10-6 to 7.8 x 10-6 at 1 µg/m3 of benzene in air), depending on
which exposure measurements were used, i.e., Crump and Allen (1984) or
Paustenbach et al. (1993). The choice of these limits was dictated by
the following considerations: (1) use of the (1981, 1987) Rinsky cohort,
(2) use of Crump's (1992, 1994) analysis of the Crump and Allen (1984)
and the Paustenbach (1992, 1993) exposure measurements. The range of
risks nearly includes the 1985 EPA risk estimate of 2.6 x 10-2 at 1 ppm
(8.1 x 10-6 at 1 µg/m3). The set of risk estimates falling within this
interval reflects both the inherent uncertainties in the risk assessment
of benzene and the limitations of the epidemiologic studies in
determining dose-response and exposure data. 

__II.C.4. Discussion of Confidence (Carcinogenicity, Inhalation
Exposure)

The major conclusion of this update (U.S. EPA, 1998) is a reaffirmation
within an order of magnitude of the benzene interim unit risk estimates
derived in EPA's 1985 interim risk assessment (U.S. EPA, 1985), which
established the probability of humans developing cancer from exposure to
1 ppm of benzene. Review of the 1985 interim risk assessment required
addressing two main concerns. 

The first concern was the use of the updated epidemiologic data from
Rinsky et al.'s (1987) cohort of Pliofilm workers and selection of
appropriate estimates of their exposure to benzene for the derivation of
the unit risk estimate. Although numerous epidemiological studies
demonstrate an association of exposure to benzene and increased risk of
human cancer, these studies are not without methodological limitations.
The Rinsky et al. (1981, 1987) study continues to provide the best
available data for derivation of unit cancer risk estimates. This study
had the least number of confounders and a wide range of exposure to
benzene. 

The second major concern was continued application of the low-dose
linearity concept to the model used to generate estimates of unit risk.
It was concluded that at present there is insufficient evidence to
reject this concept, and a linear extrapolation was used (U.S. EPA,
1998). If one assumes that the linear extrapolation model is the
appropriate model to be used, given the uncertainties outlined, then the
range of suitable estimates is defined by the choice of exposure
estimates selected. The lowest unit risk among linear choices is
determined by the exposure estimates of Paustenbach et al. (1993)
according to the calculations of Crump (1992, 1994), simply because
Paustenbach's exposure estimates for the Rinsky cohort are highest. That
estimate is 7.1 x 10-3 at 1 ppm (2.2 x 10-6 at 1 µg/m3). The highest
risk number is determined by Crump (1992, 1994) using the lower exposure
estimates from Crump and Allen (1984), and that is 2.5 x 10-2 at 1 ppm
(7.8 x 10-6 at 1 µg/m3). 

At present, the true cancer risk from exposure to benzene cannot be
ascertained, even though dose-response data are used in the quantitative
cancer risk analysis, because of uncertainties in the low-dose exposure
scenarios and lack of clear understanding of the mode of action. A range
of estimates of risk is recommended, each having equal scientific
plausibility. The range estimates are maximum likelihood values (i.e.,
best statistical estimates) and were derived from observable dose
responses using a linear extrapolation model to estimate low
environmental exposure risks. The extrapolation range is on the order of
20-60 depending on what environmental level is of interest. This range
is fairly low and thus does not suggest any unusual lack of plausibility
about the estimates. The use of a linear model is a default public
health protective approach and an argument both for and against
recognizing supra- and sublinear relationships at low doses and
nonthreshold or threshold modes of action on exposure to benzene.
Therefore, the true risk could be either higher or lower. The numerical
difference between the 1985 risk estimate (2.6 x 10-2 at 1 ppm or 8.1 x
10-6 at 1 µg/m3) compared to the new high-end risk (2.5 x 10-2 at 1 ppm
or 7.8 x 10-6 at 1 µg/m3) is insignificant and no scientific inferences
about the merit of one value versus the other should be made. 

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_II.D. EPA Documentation, Review, and Contacts (Carcinogenicity
Assessment)

__II.D.1. EPA Documentation

Source Documents -- U.S. EPA, 1999; U.S. EPA, 1998; U.S. EPA, 1985; U.S.
EPA, 1979. 

The U.S. EPA. (1998) inhalation assessment and the U.S. EPA (1999)
extrapolation of the inhalation unit risk estimate to the oral route of
exposure were externally peer reviewed. Their comments have been
evaluated carefully and incorporated in finalization of this IRIS
Summary. A summary record of the comments and EPA responses is included
as an appendix to the benzene support document file. 

The EPA 1979 and 1985 documents provide the basis for the classification
of benzene as a Group A carcinogen. 

__II.D.2. EPA Review (Carcinogenicity Assessment)

Agency Consensus Date:

inhalation carcinogenicity: 9/30/1998

oral carcinogenicity: 1/3/2000

__II.D.3. EPA Contacts (Carcinogenicity Assessment)

Please contact the IRIS Hotline for all questions concerning this
assessment or IRIS, in general, at (202)566-1676 (phone), (202)566-1749
(fax), or   HYPERLINK "mailto:hotline.iris@epa.gov" 
hotline.iris@epa.gov  (Internet address). 

_III.  [reserved]

_IV.  [reserved] 

_V.  [reserved]

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_VI.  Bibliography 

Benzene

CASRN — 71-43-2

Last Revised — 04/17/2003

_VI.A. Oral RfD References

Aksoy, M. 1989. Hematotoxicity and carcinogenicity of benzene. Environ.
Health Perspect. 82: 193-197. 

Goldstein, B.D. 1988. Benzene toxicity. Occupational medicine. State of
the Art Reviews. 3: 541-554. 

Hsieh, G.C., R.P. Sharma, and R.D.R. Parker. 1988. Subclinical effects
of groundwater contaminants. I. Alteration of humoral and cellular
immunity by benzene in CD-1 mice. Arch. Environ. Contam. Toxicol. 17:
151-158. 

NTP (National Toxicology Program). 1986. Toxicology and Carcinogenesis
Studies of Benzene (CAS No. 71-43-2) in F344/N Rats and B6C3F1 Mice
(Gavage Studies). NTP, Research Triangle Park, NC. 

Ross, D. 1996. Metabolic basis of benzene toxicity. Eur. J. Haematol.
57: 111-118. 

Rothman, N., G.L. Li, M. Dosemeci, W.E. Bechtold, G.E. Marti, Y.Z. Wang,
M. Linet, L.Q. Xi, W. Lu, M.T. Smith, N. Titenko-Holland, L.P. Zhang, W.
Blot, S.N. Yin, and R.B. Hayes. 1996. Hematotoxicity among Chinese
workers heavily exposed to benzene. Am. J. Ind. Med. 29: 236-246. 

Sabourin, P.J., B.T. Chen, G. Lucier, L.S. Birnbaum, E. Fisher, and R.F.
Henderson. 1987. Effect of dose on the absorption and excretion of
[C14]benzene administered orally or by inhalation in rats and mice.
Toxicol. Appl. Pharmacol. 87: 325-336. 

Sabourin, P.J., W.E. Bechtold, W. Griffith, L.S. Birnbaum, G. Lucier and
R.F. Henderson. 1989. Effect of exposure concentration, exposure rate,
and route of administration on metabolism of benzene by F344 rats and
B6C3F1 mice. Toxicol. Appl. Pharmacol. 99: 421-444. 

Snyder, R., G. Witz, and B.D. Goldstein. 1993. The toxicology of
benzene. Environ. Health Perspect. 100: 293-306. 

U.S. EPA (U.S. Environmental Protection Agency). 1985. Final Draft for
Drinking Water Criteria Document on Benzene. Office of Drinking Water,
Washington, DC. PB86-118122. 

U.S. EPA. 1989. Workgroup for Risk Assessment Guidance for Superfund.
Volume 1. Human Health Evaluation Manual. Part A. Office of Solid Waste
and Emergency Response, Washington, DC. EPA/540/1-89/002. 

U.S. EPA. 1994. Methods for derivation of inhalation reference
concentrations and application of inhalation dosimetry,
EPA/600/8-90/066F, dated October, 1994. 

U.S. EPA. 1999. Extrapolation of the Benzene Inhalation Unit Risk
Estimate to the Oral Route of Exposure. National Center for
Environmental Assessment, Office of Research and Development,
Washington, DC. NCEA-W-0517. 

U.S. EPA. 2000. Benchmark Dose Technical Guidance Document (External
Review Draft). EPA/630/R-00/001. 

U.S. EPA. 2002. Toxicological Review of Benzene (Noncancer Effects).
Available online at:   HYPERLINK "http://www.epa.gov/ncea/iris" 
www.epa.gov/iris . 

White, K.L. Jr., H.H. Lysy, J.A. Munson, et al. 1984. Immunosuppression
of B6C3F1 female mice following subchronic exposure to benzene from
drinking water. TSCA 8E Submission. OTS Fiche # OTS0536214.

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_VI.B. Inhalation RfC References

Aksoy, M. 1989. Hematotoxicity and carcinogenicity of benzene. Environ.
Health Perspect. 82: 193-197. 

Aksoy, M., K. Dincol, K. Erdem, T. Akgun, and G. Dincol. 1972. Details
of blood changes in 32 patients with pancytopenia associated with
long-term exposure to benzene. Br. J. Ind. Med. 29: 56-64. 

ATSDR (Agency for Toxic Substances and Disease Registry) 1997.
Toxicological profile for benzene (Update). Public Health Service, U.S.
Department of Health and Human Services, Atlanta, GA. 

Baarson, K.A., C.A. Snyder, and R.E. Albert. 1984. Repeated exposure of
C57B1 mice to inhaled benzene at 10 ppm markedly depressed
erythropoietic colony formation. Toxicol. Lett. 20: 337-342. 

Bogardi-Sare, A., M. Zavalic, I. Trosic et al. 2000. Study of some
immunological parameters in workers occupationally exposed to benzene.
Int. Arch. Occup. Environ. Health. 73: 397-400. 

Cronkite, E.P., R.T. Drew, T. Inoue and J.E. Bullis. 1985. Benzene
hematotoxicity and leukemogenesis. Am. J. Ind. Med. 7: 447-456. 

Dosemeci, M., S-N. Yin, M. Linet et al. 1996. Indirect validation of
benzene exposure assessment by association with benzene poisoning.
Environ. Health Perspect. 104(Suppl. 6): 1343-1347. 

Goldstein, B.D. 1988. Benzene toxicity. Occupational medicine. State of
the Art Reviews. 3: 541-554. 

Keller, K.A. and C.A. Snyder. 1986. Mice exposed in utero to low
concentrations of benzene exhibit enduring changes in their colony
forming hematopoietic cells. Toxicology. 42: 171-181. 

Keller, K.A. and C.A. Snyder. 1988. Mice exposed in utero to 20 ppm
benzene exhibit altered numbers of recognizable hematopoietic cells up
to seven weeks after exposure. Fund. Appl. Toxicol. 10: 224-232. 

Ross, D. 1996. Metabolic basis of benzene toxicity. Eur. J. Haematol.
57: 111-118. 

Rothman, N., G.L. Li, M. Dosemeci, W.E. Bechtold, G.E. Marti, Y.Z. Wang,
M. Linet, L.Q. Xi, W. Lu, M.T. Smith, N. Titenko-Holland, L.P. Zhang, W.
Blot, S.N. Yin, and R.B. Hayes. 1996. Hematotoxicity among Chinese
workers heavily exposed to benzene. Am. J. Ind. Med. 29: 236-246. 

Snyder, C.A., B.D. Goldstein, A.R. Sellakumar, I. Bromberg, S. Laskin,
and R.E. Albert. 1980. The inhalation toxicity of benzene: Incidence of
hematopoietic neoplasms and hematoxicity in AKR/J and C57BL/6J mice.
Toxicol. Appl. Pharmacol. 54: 323-331. 

Snyder, R., G. Witz, and B.D. Goldstein. 1993. The toxicology of
benzene. Environ. Health Perspect. 100: 293-306. 

U.S. EPA (U.S. Environmental Protection Agency). 1994. Methods for
Derivation of Inhalation Reference Concentrations and Application of
Inhalation Dosimetry. Prepared by the Office of Health and Environmental
Assessment, Research Triangle Park, NC. EPA/600/8-90/066F. 

U.S. EPA. 2000. Benchmark Dose Technical Guidance Document (External
Review Draft). EPA/630/R-00/001. 

U.S. EPA. 2002. Toxicological Review of Benzene (Noncancer Effects).
Available online at:   HYPERLINK "http://www.epa.gov/ncea/iris" 
www.epa.gov/iris . 

Ward, E., R. Hornung, J. Morris, R. Risnsky, D. Wild, W. Halperin, and
W. Guthrie. 1996. Risk of low red or white blood cell count related to
estimated benzene exposure in a rubberworker cohort (1940-1975). Am. J.
Ind. Med. 29: 247-257. 

Ward, C.O., R.A. Kuna, N.K. Snyder, R.D. Alsaker, W.B. Coate, and P.H.
Craig. 1985. Subchronic inhalation toxicity of benzene in rats and mice.
Am. J. Ind. Med. 7: 457-473.

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_VI.C. Carcinogenicity Assessment References

ATSDR (Agency for Toxic Substances and Disease Registry). (1997)
Toxicological profile for benzene. Update. Public Health Service, U.S.
Department of Health and Human Services, Atlanta, Ga. 

Bois, FY; Jackson, ET; Pekari, K; et al. (1996) Population
toxicokinetics of benzene. Environ Health Perspect 104 (Suppl
6):1405-1411. 

Crump, KS. (1994) Risk of benzene-induced leukemia: a sensitivity
analysis of the Pliofilm cohort with additional follow-up and new
exposure estimates. J Toxicol Environ Health 42:219-242. 

Crump, KS; Allen, BC. (1984) Quantitative estimates of risk of leukemia
from occupational exposure to benzene. Prepared for the Occupational
Safety and Health Administration by Science Research Systems, Inc.,
Ruston, LA. Unpublished 

Fiserova-Bergerova, V; Vlach, J; Singhal, K. (1974) Simulation and
prediction of uptake, distribution, and exhalation of organic solvents.
Br J Ind Med 31:45-52. 

Gerrity, TR Henry, CJ, eds. (1990) Principles of route-to-route
extrapolation for risk assessment: Proceedings of the Workshops on
Principles of Route-to-route Extrapolation for Risk Assessment, held
1990: Hilton Head, SC, and Durham, NC. New York: Elsevier. 

Hunter, CG. (1966) Aromatic solvents. Ann Occup Hyg 9:191-198. 

Hunter, CG. (1968) Solvents with reference to studies on the
pharmacodynamics of benzene. Proc R Soc Med 61:913-915. 

Hunter, CG; Blair, D. (1972) Benzene: pharmacokinetic studies in man.
Ann Occup Hyg 15:193-201. 

Nomiyama, K; Nomiyama, H. (1974) Respiratory retention, uptake and
excretion of organic solvents in man. Int Arch Arbeitsmed 32:75-83. 

Paustenbach, D; Bass, R; Price, P. (1993) Benzene toxicity and risk
assessment, 1972-1992: implications for future regulation. Environ
Health Perspect 101 (Suppl 6):177-200. 

Pekari, K; Vainiotalo, S; Heikkila, P; et al. (1992) Biological
monitoring of occupational exposure to low levels of benzene. Scand J
Work Environ Health 18:317-322. 

Rinsky, RA; Young, RJ; Smith, AB. (1981) Leukemia in benzene workers. Am
J Ind Med 2:217-245. 

Rinsky, RA; Smith, AB; Horning, R; et al. (1987) Benzene and leukemia:
an epidemiologic risk assessment. N Engl J Med 316:1044-1050. 

Sabourin, PJ; Chen, BT; Lucier, G; et al. (1987) Effect of dose on the
absorption and excretion of [14C] benzene administered orally or by
inhalation in rats and mice. Toxicol Appl Pharmacol 87:325-336. 

Sherwood, RJ. (1988) Pharmacokinetics of benzene in a human after
exposure at about the permissible limit. Ann N Y Acad Sci 534:635-647. 

Smith, MT. (1996) The mechanism of benzene-induced leukemia: a
hypothesis and speculations on the causes of leukemia. Environ Health
Perspect 104 (Suppl 6):1219-1225. 

Smith, MT; Fanning, EW. (1997) Report on the workshop entitled:
"Modeling chemically induced leukemia-implications for benzene risk
assessment." Leuk Res 21:361-374. 

Srbova, J; Teisinger, J; Skramovsky, S. (1950) Absorption and
elimination of inhaled benzene in man. Arch Ind Hyg Occup Med 2:1-8. 

U.S. EPA. (1992, April 1) Integrated Risk Information System (IRIS).
Substance file - benzene. Washington, DC: National Center for
Environmental Assessment. 

U.S. EPA. (1998, April 10) Carcinogenic effects of benzene: an update.
Prepared by the National Center for Environmental Health, Office of
Research and Development. Washington, DC. EPA/600/P-97/001F. 

U.S. EPA. (1999) Extrapolation of the benzene inhalation unit risk
estimate to the oral route of exposure. National Center for
Environmental Health, Office of Research and Development. Washington,
DC. NCEA-W-0517. 

Yu, R; Weisel, CP. (1998) Measurement of benzene in human breath
associated with an environmental exposure. J Expo Anal Environ Epidemiol
6,3:261-277. 

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_VII.  Revision History

Benzene

CASRN — 71-43-2

Date	Section	Description

12/01/1988	II.A.4.	Anderson and Richardson citation year corrected

12/01/1988	II.A.4.	Kissling and Speck citation year corrected

07/01/1989	I.B.	Inhalation RfD now under review

02/01/1990	II.	Clarified citations

02/01/1990	II.A.3.	Corrected Maltoni, 1979 to Maltoni and Scarnato, 1979

02/01/1990	II.A.3.	Corrected Maltoni, 1983 to Maltoni et al., 1983

02/01/1990	II.A.3.	Corrected Synder et al., 1980 to 1981

02/01/1990	VI.	Bibliography on-line

03/01/1990	VI.C.	Clarify Maltoni et al., 1983 and NTP, 1986 references 

08/01/1990	III.A.10	Primary contact changed

08/01/1990	IV.F.1.	EPA contact changed

01/01/1991	II.	Text edited

01/01/1991	II.C.1.	Inhalation slope factor removed (global change)

01/01/1992	IV.	Regulatory actions updated

04/01/1992	II.B.2.	Text revised

02/01/1994	II.D.3.	Secondary contact's phone number changed

08/01/1995 	I.B.	EPA's RfD/RfC and CRAVE workgroups were discontinued in
May, 1995. Chemical substance reviews that were not completed by
September 1995 were taken out of IRIS review. The IRIS Pilot Program
replaced the workgroup functions beginning in September, 1995.

04/01/1997	III., IV., V.	Drinking Water Health Advisories, EPA
Regulatory Actions, and Supplementary Data were removed from IRIS on or
before April 1997. IRIS users were directed to the appropriate EPA
Program Offices for this information.

10/16/1998	II., VI.	Revised inhalation carcinogenicity section and
references

01/19/2000 	II., VI.	Revised oral carcinogenicity section and references

02/22/2001	I.	This chemical is being reassessed under the IRIS Program.

04/17/2003	IA., IB., VI.	New RfD and RfC sections; references.

10/01/2008	I.A.2.	Text change made to the first sentence under the Table
1 footnote.



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_VIII.  Synonyms

Benzene

CASRN — 71-43-2

Last Revised — 01/19/2000

71-43-2 

Benzene 

benzol 

coal naphtha 

cyclohexatriene 

phene 

phenyl hydride 

polystream 

pyrobenzol

