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Posted Date: 2010-11-30T05:00Z

1,3-Butadiene (CASRN 106-99-0) 

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0139

1,3-Butadiene; CASRN 106-99-0; 11/05/2002

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 1,3-Butadiene

File First On-Line 03/31/1987

Category (section)	Status	Last Revised

Oral RfD Assessment (I.A.)	no data	 

Inhalation RfC Assessment (I.B.)	on-line	11/05/2002

Carcinogenicity Assessment (II.)	on-line	11/05/2002

_I.  Chronic Health Hazard Assessments for Noncarcinogenic Effects

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

1,3-Butadiene

CASRN — 106-99-0

Last Revised — 11/05/2002 

The oral Reference Dose (RfD) is based on the assumption that thresholds
exist for certain toxic effects such as hemolysis. 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 Background Documents for an elaboration of these concepts.
RfDs can also be derived for the noncarcinogenic health effects of
substances that are also 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.

An oral RfD is not calculated because 1,3-butadiene is a gas and causes
hazard by inhalation only. The hazard by ingestion is unlikely since
1,3-butadiene is poorly soluble in water. When released in water,
1,3-butadiene rapidly evaporates.

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

Substance Name — 1,3-Butadiene

CASRN — 106-99-0

Last Revised — 11/05/2002 

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 systems peripheral to the respiratory system. It
is generally expressed in units of mg/m3. 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 the Interim Methods for Development of Inhalation Reference
Doses (EPA/600/8-88/066F, August 1989) and subsequently, according to
Methods for Derivation of Inhalation Reference Concentrations and
Application of Inhalation Dosimetry (EPA/600/8-90/066F, October 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

A variety of reproductive and developmental effects have been observed
in mice exposed to 1,3-butadiene by inhalation (U.S. EPA, 2002, Chapter
5). There are no human data on reproductive or developmental effects.
Few adverse noncancer effects, other than reproductive and developmental
effects, have been observed, except for hematological effects in mice
exposed to higher concentrations (U.S. EPA, 2002, Section 6.1).

The most sensitive short-term developmental endpoint was decreased fetal
weight in the mouse. Decreases were observed at the lowest exposure
concentration (40 ppm, 6 hours/day, gestation days 6-15); thus there was
not a no-observed-adverse-effect level (NOAEL) for this effect (Hacket
et al., 1987). No developmental toxicity was observed in rats. 

The most sensitive reproductive endpoint observed in subchronic exposure
studies was fetal deaths in dominant lethal studies of mice (i.e., male
mice were exposed to 1,3-butadiene and effects on litters were measured
after mating to unexposed females) (Anderson et al., 1998; Brinkworth et
al., 1998; Anderson et al., 1993; Adler and Anderson, 1994). Significant
dominant lethal effects were observed at exposures of 65 ppm, 6
hours/day, 5 days/week, for 4 weeks. (The 12.5 ppm exposure level was a
NOAEL.) Dominant lethal effects in humans would likely be manifested as
infertility (due to reduced fertility or very early deaths) or
spontaneous abortions. The dominant lethal responses are believed to
represent a genotoxic effect. 

From chronic exposure studies (2-year bioassays; NTP, 1993), the most
sensitive reproductive effects were ovarian atrophy in female mice and
testicular atrophy in male mice. Testicular atrophy was primarily a
high-exposure effect. Ovarian atrophy, on the other hand, was observed
at the lowest exposure level (6.25 ppm, 6 hours/day, 5 days/week, for 2
years). Uterine atrophy was also observed in the highest exposure
groups; however, this is likely to be a secondary effect of the ovarian
atrophy. The mechanisms of ovarian atrophy are unknown, although there
is strong evidence that the effect is mediated by the diepoxide
metabolite (U.S. EPA, 2002, Chapter 5).

Critical Effect	Experimental Doses*	UF	MF	RfC

Ovarian atrophy

2-year mouse

inhalation study

(NTP, 1993)	BMCL10 = 0.88 ppm (HEC)

(BMC10 = 1.0 ppm)	1000	1	0.9 ppb

(2 × 10-3; mg/m3)

*Conversion Factors and Assumptions — ppm equivalence across species
was assumed (this is the same as using EPA's inhalation dosimetry
methodology with RGDRr=1 [U.S. EPA, 1994]); exposure concentrations were
adjusted to 24-hour continuous daily exposure for the exposure period
(i.e., exposure concentration × [6/24] × [5/7]). 1 ppm = 2.25 mg/m3.

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

The chronic RfC of 0.9 ppb is based on ovarian atrophy. A BMCL10 (0.88
ppm) was calculated from data from the 1993 NTP 2-year inhalation
bioassay, including interim kill data, using benchmark concentration
methodology (Weibull time-to-response model). In this bioassay, groups
of 70 female B6C3F1 mice were exposed by inhalation 6 hours/day, 5
days/week to 0, 6.25, 20, 62.5, or 200 ppm 1,3-butadiene for up to 103
weeks; groups of 90 female mice were exposed to 625 ppm. Interim
evaluations were conducted at 9 and 15 months on up to 10 mice per
group. Significant concentration-related decreases in survival were seen
in female mice exposed to concentrations >= 20 ppm, primarily due to the
development of malignant neoplasms. Statistically significant increases
in the incidence of ovarian atrophy were observed in all exposure
groups, including the lowest exposure group (6.25 ppm), following
lifetime exposures. In calculating the BMC10 and BMCL10, lesion severity
was not taken into account, and the 625 ppm group was excluded because
of high early mortality. In addition, ovarian atrophy was modeled to
reflect extra risks only until age 50, because 1,3-butadiene-induced
ovarian atrophy is believed to result from follicular failure, and after
menopause, follicles would no longer be available. Exposure
concentrations were converted to 24-hour exposures by multiplying by
(6/24) and (5/7). 

Benchmark dose modeling and sample RfC calculations were also conducted
for the endpoints of fetal weight (7 ppb), dominant lethal effects (20
ppb), and testicular atrophy (20 ppb) (U.S. EPA, 2002, Sections 10.3 and
10.4). Ovarian atrophy was selected as the critical effect because it
yielded the lowest RfC. Ovarian atrophy also had the lowest BMC10 and
was reported in a high-quality 2-year study.

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

UF = 1000. 

For ovarian atrophy, the uncertainty/modifying factors were: 3 for
interspecies extrapolation, 10 for intraspecies variability, 3 for
incomplete database, and 10 for extrapolation to a level below the 10%
effect level (analogous to the LOAEL-to-NOAEL extrapolation factor). The
BMCL10 was estimated from a chronic bioassay; therefore, an
acute/subchronic-to-chronic factor was not required. The factor of 10
used for effect level extrapolation was derived from a formula¹ that
takes into account the benchmark response level as well as the slope of
the exposure-response model at the benchmark concentration. However,
because the model was supralinear at the BMC10, a maximum factor of 10
for the 10% response level was used (U.S. EPA, 2002, Chapter 10). EPA is
planning to develop guidance for applying an effect level extrapolation
factor to a benchmark dose; the formula mentioned above was used in the
interim. An extrapolation factor for effect level is applied because the
10% response level used as the point of departure is an adverse effect
level. Therefore, a factor analogous to the LOAEL-to-NOAEL factor is
needed to attempt to extrapolate to a level closer to a no effect level.
The default factors of 3 for interspecies extrapolation for inhalation
exposures and of 10 for intraspecies variability were used. There is
strong evidence that the diepoxide metabolite (1,2:3,4-diepoxybutane,
DEB) is required to elicit ovarian atrophy (U.S. EPA, 2002, Chapter 5),
and it is expected, based on pharmacokinetic data, that humans produce
less DEB than mice (U.S. EPA, 2002, Chapter 3). However, DEB levels
cannot be quantified without an adequate physiologically based
pharmacokinetic (PBPK) model. Thus, default dosimetry (i.e.,
1,3-butadiene exposure concentration) was used for dose-response
modeling, and the default value of 1 for the pharmacokinetic portion of
the interspecies uncertainty factor for inhalation exposures was
retained. Finally, a factor of 3 was used to reflect an incomplete
database, in particular the absence of a multigeneration study and a
developmental neurotoxicity study. Dividing the BMCL10 of 0.88 ppm by
the composite UF of 1000 yields 0.9 ppb. 

MF = 1.

¹ The formula is as follows: uncertainty factor = x × [(slope of the
line from the BMCx to 0)/(slope of the dose-response curve at the
BMCx)], where x% is the response level. Results of the formula are
confined within a minimum value of 3 and a maximum value of x.

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

A subchronic inhalation study showed that just 13 weeks of exposure to
1,000 ppm 1,3-butadiene was also sufficient to induce ovarian atrophy in
female B6C3F1 mice (Bevan et al., 1996). 

In addition to deriving the chronic RfC, the Health Assessment Document
also provides reference concentration values for acute and subchronic
exposure scenarios, based on the mouse fetal weight data of Hackett et
al. (1987) (see U.S. EPA, 2002, Chapter 10, Sections 10.3.2 and 10.4).
These reference concentration values are not currently part of the IRIS
file. 

The EPA closely examined the physiologically-based pharmacokinetic
(PBPK) models for 1,3-butadiene to determine if additional modeling
could reduce uncertainties in the interspecies scaling between mice and
humans for ovarian atrophy and other endpoints (U.S. EPA, 2002, Chapter
9). Despite advances in the models over the past decade, the current
models are inadequate for this purpose. For example, the PBPK models do
not yet accurately describe the distribution of the major metabolites in
various compartments, they do not yet include the reportedly important
epoxydiol metabolite, and they have not been adequately validated.

__I.B.5. Confidence in the Inhalation RfC

Study — High

Database — High

RfC — Medium 

The overall confidence in this RfC assessment is medium. The RfC
calculation was based on data from a high-quality NTP 2-year bioassay in
which many exposure levels were used, although a NOAEL was not achieved.
On the other hand, rat studies showed no such evidence of reproductive
and developmental effects, and there are no human data on these effects;
thus, it is uncertain how humans would respond. 

__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 (the Science
Advisory Board). Their comments have been evaluated carefully and
incorporated in finalization of this IRIS Summary. A record of these
comments is included as an appendix to U.S. EPA, 2002. 

Other documentation -- U.S. EPA, 1985

Agency Consensus Date — 9/13/2001

__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  (email address).

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_II.  Carcinogenicity Assessment for Lifetime Exposure

Substance Name — 1,3-Butadiene

CASRN — 106-99-0

Last Revised — 11/05/2002

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
µg/L drinking water or risk per µg/m3 air breathed. The third form in
which risk is presented is a concentration of the chemical in drinking
water or air associated with 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
Documents. 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

This weight-of-evidence carcinogenicity classification and quantitative
estimate of carcinogenicity from inhalation exposure replace the
previous classification of "B2; probable human carcinogen," and
inhalation unit risk of 2.8 × 10-4 per µg/m3, entered on IRIS on
March, 31, 1987. The new classification and unit risk estimate are based
on more recent data. 

Under EPA's 1999 Guidelines for Carcinogen Risk Assessment (U.S. EPA,
1999), 1,3-butadiene is characterized as carcinogenic to humans by
inhalation. This characterization is supported by the total
weight-of-evidence provided by the following: (1) sufficient evidence
from epidemiologic studies of the majority of U.S. workers
occupationally exposed to 1,3-butadiene, either to the monomer or to the
polymer by inhalation, showing increased lymphohematopoietic cancers and
a dose-response relationship for leukemias in polymer workers (see
Section II.A.2), (2) sufficient evidence in laboratory animal studies
showing that 1,3-butadiene causes tumors at multiple sites in mice and
rats by inhalation (see Section II.A.3), and (3) numerous studies
consistently demonstrating that 1,3-butadiene is metabolized into
genotoxic metabolites by experimental animals and humans (see Section
II.A.4). The specific mechanisms of 1,3-butadiene-induced carcinogenesis
are unknown; however, the scientific evidence strongly suggests that the
carcinogenic effects are mediated by genotoxic metabolites of
1,3-butadiene, i.e., the monoepoxide, the diepoxide, and the epoxydiol. 

__II.A.2. Human Carcinogenicity Data

There is "sufficient evidence" from epidemiologic studies of exposed
workers to consider 1,3-butadiene carcinogenic to humans. The exposure
to 1,3-butadiene occurs in monomer production workers who produce
1,3-butadiene as a raw material or in polymer production workers who use
1,3-butadiene in styrene-butadiene rubber (SBR) production. Excesses of
lymphohematopoietic cancers have been observed in 1,3-butadiene polymer
production workers and monomer production workers in North America. A
significant excess of leukemias was observed in polymer production
workers, and significant excesses of non-Hodgkin's lymphomas (previously
diagnosed as lymphosarcoma and reticular sarcoma, but now included in
non-Hodgkin's lymphomas [NHL] per the new classification in the
International Classification of Diseases of Oncology [ICD-O]) have been
observed in monomer workers. Under the previous, as well as the current,
classification system adopted by the Revised European-American Lymphoma
(REAL) and the Leukemia Society of America², both leukemia and lymphoma
are lymphohematopoietic cancers, and thus the lymphohematopoietic system
is considered to be the target organ for 1,3-butadiene. 

The strongest evidence is provided by a retrospective cohort study of
over 15,000 SBR workers in 8 plants studied by the University of Alabama
at Birmingham (UAB cohort), with 49 years of follow-up (Delzell et al.,
1996). Quantitative exposures (cumulative and peak) to 1,3-butadiene,
styrene, and benzene were estimated for each worker (Macaluso et al.,
1996). Limited validations of exposure estimates were attempted by
various means. Significant excesses ranging from 43% to 336% were found
for leukemia in ever hourly workers as compared with the general
population, after adjusting for styrene and benzene. An internal
comparison, using estimated ppm-years of 1,3-butadiene exposure,
resulted in increasing risk ratios for leukemia with increasing
exposures. This trend was statistically significant. A fairly consistent
association between exposure to 1,3-butadiene and occurrence of leukemia
across the six plants for which subanalyses were done was also found. 

The major strengths of this study are the detailed and comprehensive
quantitative exposure estimations for 1,3-butadiene, styrene, and
benzene for each individual. The cohort was also large, and there was a
long follow-up period of 49 years. In addition, both external and
internal comparison showed similar results, adjustments for potential
confounding factors were carried out, and analyses by duration of
employment and for latency were conducted. 

The study had some limitations. Some misclassifications of exposure may
have occurred with respect to certain jobs, but these are unlikely to
have occurred only in leukemia cases because the exposures were
calculated a priori. Furthermore, the excess mortality observed for
leukemia was based on death certificates and was not verified by medical
records, thus, there may be misclassification of diagnosis. The
histologic typing of leukemia was also not available, so currently it is
not known whether a single cell type or more than one cell type is
associated with the exposure to 1,3-butadiene. Two plants were
eliminated from the final analysis due to the lack of work histories,
which may have resulted in the loss of valuable data. Finally, an issue
has been raised of potential confounding exposure to dithiocarbamates
(DTC) (Irons and Pyatt, 1998). DTCs have been in use since the early
1940s as fungicides and treatments for parasitic skin diseases. The DTC
disulfiram has also been in use since the early 1940s for the treatment
of alcoholism. So far, there is not even a case report of leukemia in
the literature in reference to any of the DTCs. In addition, available
animal studies have not provided any evidence that DTCs cause
carcinogenesis. Hence, at this time, it is conjecture that DTCs are
causally associated with leukemia and, therefore, confound the results
of Delzell et al. 

Additional evidence is provided by the earlier cohort study of some of
these polymer workers (13,500 individuals; Johns Hopkins University
[JHU] cohort)3 conducted by Matanoski and Schwartz (1987), as well as a
nested case-control study by Santos-Burgoa et al. (1992). A significant
excess of lymphohematopoietic cancer, but not of leukemia, was observed
in the cohort study, while a significantly increased odds ratio of 7 for
leukemia was observed in a nested case-control study, as was a
significant trend of increasing risk of leukemia with increasing
exposure level of 1,3-butadiene. 

For 1,3-butadiene monomer production workers, two of three different
cohort studies found significant increased risk of NHL (previously
classified as lymphosarcoma and reticulosarcoma) (Divine and Hartman,
1996; Ward et al., 1995). The third study (Cowles et al., 1994) was
small and had shorter follow-up times. The strongest evidence of human
carcinogenicity from monomer production worker studies is provided by
the largest cohort of approximately 2,800 workers in a Texaco plant
studied by several investigators (Downs et al., 1987; Divine and
Hartman, 1996). The only significant excess mortality observed was for
lymphosarcoma (now included in NHL) in the wartime subcohort of workers
(154% to 169% higher than the general population). The investigators
estimated exposures for each individual in their last follow-up (Divine
and Hartman, 1996) and found that, except for an excess observed for NHL
(76% higher than the general population) in the wartime subcohort, there
were no excesses in any cause-specific cancer mortality. 

The major strengths of this study are a relatively large cohort of
monomer workers, a long follow-up period of 52 years, analyses by
duration of employment and latency, and adjustment for potential
confounding factors. The exposures in each individual were estimated in
the last follow-up. Except for "hire-age"4 in survival analysis using
the Cox model, after 52 years of follow-up, this study did not find any
statistically significant excess in leukemia (as was observed in polymer
workers), although an increase of 13% was reported. This study may not
have enough statistical power to detect a significant leukemia increase.

Some of the limitations of the study are a lack of data or means
available to the investigators to estimate the peak exposures that were
hypothesized to be associated with the observed increase in
lymphosarcomas in wartime workers. The authors' claim of the existence
of extremely high peak exposures during the 1950s and 1960s cannot be
validated in the absence of any information about the frequency or the
variations in intensity of peak exposures for these different time
periods (as compared to prior to the 1950s). Although the cohort is
relatively large, it had low power to detect excess leukemias.
Nonetheless, the finding of excess mortality from lymphosarcoma is
consistent with the findings of Ward et al. (1995). 

Ward et al. (1995) studied a small cohort of 364 individuals who had
potential exposure to 1,3-butadiene at three Union Carbide plants. A
statistically significant excess for lymphosarcoma (477% higher than the
general population) was found based on 4 cases. The main limitations of
this study are that the cohort was small and that exposures were assumed
based on department codes. In addition, there was no analysis for
latency or adjustment for potential confounding by exposure to other
chemicals. 

These monomer and polymer production worker cohorts demonstrate an
excess number of lymphohematopoietic cancers in occupationally exposed
workers. Increased NHL (lymphosarcomas) are reported for monomer
production workers, whereas excess leukemias occur predominantly in
polymer production workers. There are several possible explanations for
this apparent difference between the monomer and polymer workers. It has
been hypothesized that the observed excess of NHL (lymphosarcomas) in
the monomer production workers may be related to exposure intensity,
i.e., the excess risk may result from the high (peak) exposures during
wartime, rather than the much lower exposures currently encountered by
monomer production workers or the likewise comparatively lower exposures
encountered by the polymer production workers. The absence of a
significant leukemia excess in these same monomer workers may be
attributable to low statistical power in the monomer studies. There is
some suggestion of excess leukemias in the monomer production workers,
although these were not statistically significant. The Union Carbide
cohort had a leukemia excess of 23% based on 2 cases, and the Texaco
cohort had an elevated risk of leukemias of 13% based on 13 cases (it
should be noted though, that with every follow-up of the Texaco cohort,
investigators have observed additional leukemia deaths). Even the Texaco
cohort, a relatively large monomer production cohort, has low power to
detect a statistically significant excess for leukemias, and with every
follow-up, the investigators of the Texaco cohort increased the calendar
period for the worker inclusion criteria. This added many younger
workers with little cumulative exposure, shorter follow-up periods, and
inadequate latency periods, thereby diluting the risk. In addition,
1,3-butadiene is produced at the end of the monomer production process,
and current 1,3-butadiene exposures are very low in these workers. 

In fact, the apparent difference between monomer and polymer workers may
be largely an artifact. Under the latest classification system for
lymphohematopoietic cancers, all lymphomas not classified as Hodgkin's
disease are now included under NHL (see footnote 1). Using this
classification, an excess of NHL of 37% (based on 15 cases; not
statistically significant) was reported for workers who had worked >= 10
years and with >= 20 years since hire in the UAB (polymer) cohort
(Sathiakumar et al., 1998). (Previously lymphosarcomas and NHL were
reported separately for this cohort.) Furthermore, as these
investigators report, their evaluation of NHL relations was limited by
their reliance on death certificates. NHL has high survival rates and
may, in later clinical stages, transform into leukemia. Therefore,
leukemia may be reported on the death certificates. In addition, as
discussed above, nonsignificant excesses of leukemia were observed in
two monomer studies. Thus, excesses of both leukemia and NHL have been
observed for both monomer and polymer workers, and it may be that the
increased risk of NHL is primarily observed among workers exposed to
high concentrations of 1,3-butadiene (mostly wartime monomer workers),
whereas the polymer production studies have greater power to detect a
significant leukemia excess among SBR workers who have modest to low
exposures. In any event, leukemias and NHL are related tumor types and
can both be classified as lymphohematopoietic cancers (see footnote 1). 

Finally, an alternate explanation is that the monomer workers may lack
exposure to a necessary co- or modifying factor that may be present in
polymer production, resulting in the development of leukemias, although
the findings of Delzell et al. (1996) and Macaluso et al. (1996) show no
evidence of confounding by exposure to other chemicals. 

In summary, the findings of excess lymphohematopoietic cancers in
polymer and monomer production workers are consistent with a causal
association with exposure to 1,3-butadiene. As demonstrated above, the
causality criteria of temporality, strength of association, specificity,
biological gradient, and consistency are satisfied. In addition, as
discussed in the next sections, 1,3-butadiene is metabolized by humans
and other species to genotoxic metabolites and is carcinogenic in mice
and rats, thus fulfilling the criterion of biological plausibility as
well. Therefore, the human evidence is considered sufficient.

² Under the previous classification (8th ICD, Adapted),
lymphohematopoietic cancers comprised the following subcategories:
lymphosarcoma and reticular sarcoma, Hodgkin's disease, leukemia, and
other lymphatic tissue cancers. In 1994, the International Lymphoma
Group's Revised European-American Lymphoma (REAL) classification was
proposed for the lymphohematopoietic cancers, and is being adopted into
the ICD-O (Berard and Hutchison, 1997). This classification is based on
new ideas evolving in the fields of molecular biology, genetics, and
immunology, which have rendered the old classification for
lymphohematopoietic cancers obsolete. The REAL classification comprises
the following subcategories: B-cell neoplasms, T-cell and putative
natural killer (NK)-cell neoplasms, Hodgkin's disease, and unclassified
lymphomas. It should be noted that both leukemias and lymphomas that are
produced by B-cells are included under B-cell neoplasms, and leukemias
and lymphomas produced by T-cells and NK-cells are included under T-cell
and NK-cell neoplasms. Any lymphoma (such as B-cell, T-cell, and
NK-cell) that is not classified as Hodgkin's disease is included under
non-Hodgkin's lymphoma. 

Furthermore, the Leukemia Society of America defines lymphohematopoietic
cancers as follows: "Leukemia, Lymphoma, Hodgkin's disease, and Myeloma
are cancers of the body's blood forming and immune systems: the bone
marrow and lymph nodes. They are considered to be related cancers
because they involve the uncontrolled growth of cells with similar
functions."

3 One Canadian plant and six U.S. plants were common in the JHU and the
UAB cohorts.

4 Survival analyses were conducted by the investigators using three
different methods in their last follow-up. Two different risk factors
were used for these analyses ([1] Exposure, i.e., cumulative exposure,
and [2] Hire-age, i.e., age at which the employee was hired) to
calculate risks for all lymphohematopoietic cancer, leukemia,
lymphosarcoma, NHL, and multiple myeloma.

__II.A.3. Animal Carcinogenicity Data

Sufficient. Several chronic inhalation bioassay studies have been
conducted with 1,3-butadiene: a 2-year rat study (Hazleton Laboratories
Ltd., 1981; Owen et al., 1987); two lifetime mouse studies (NTP, 1984,
1993), the first terminated early because of excessive mortality and the
second using lower exposure concentrations; a 2-year stop-exposure study
with male mice (NTP, 1993); and a 1-year study comparing the induction
of thymic lymphomas in two different strains of male mice (Irons et al.,
1989). These studies provide unequivocal evidence that 1,3-butadiene is
a multisite carcinogen in both rats and mice, with the mouse being
significantly more sensitive than the rat. 

In the most recent mouse study (NTP, 1993), groups of 70 male and 70
female B6C3F1 mice were exposed by inhalation 6 hours/day, 5 days/week
to 0, 6.25, 20, 62.5, or 200 ppm 1,3-butadiene for up to 103 weeks,
while groups of 90 male and 90 female mice were exposed to 625 ppm.
Interim evaluations were conducted at 9 and 15 months on up to 10 mice
per group. Significant concentration-related decreases in survival were
seen in male and female mice exposed to concentrations >= 20 ppm,
primarily due to the development of malignant neoplasms. Significant
concentration-related increases in survival-adjusted incidences were
observed for the following primary neoplasms in both males and females:
malignant lymphomas; histiocytic sarcomas; heart hemangiosarcomas;
alveolar/bronchiolar adenoma, carcinoma, or adenocarcinomas; and
forestomach squamous cell papilloma or carcinomas. Female mice also
exhibited significant concentration-related increases in
survival-adjusted incidences of benign or malignant granulosa cell
tumors (ovary) and of adenocanthoma, carcinoma, or malignant mixed
tumors of the mammary gland. Other tumor types that showed significant
increases in some exposure groups versus controls for male and/or female
mice were hepatocellular adenoma or carcinomas, Harderian gland adenoma
or carcinomas, and preputial gland carcinomas. The most sensitive site
was the female mouse lung, which exhibited significantly increased tumor
incidence at the lowest exposure concentration tested (6.25 ppm). 

In the sole rat study (Hazleton Laboratories Ltd., 1981), Charles River
CD rats (110/sex/group) were exposed by inhalation to 0, 1,000, or 8,000
ppm 1,3-butadiene 6 hours/day, 5 days/week for up to 111 weeks. There
was a treatment-related increase in mortality, some of which was
attributed to nephropathies in males. In exposed females, significant
increases occurred in incidences of mammary gland carcinoma or
fibroadenomas and thyroid follicular cell adenoma or carcinomas. In
exposed males, there were also significant increases in thyroid
follicular cell tumors, as well as in Leydig cell tumors and pancreatic
exocrine adenomas. Although not significant by pairwise comparisons,
significant exposure-response trends were observed for Zymbal gland
carcinomas and uterine stromal sarcomas in females and for brain gliomas
in males.

__II.A.4. Supporting Data for Carcinogenicity 

1,3-Butadiene is metabolized to at least three genotoxic metabolites: a
monoepoxide (1,2-epoxy-3-butene, EB), a diepoxide
(1,2:3,4-diepoxybutane, DEB), and an epoxydiol
(3,4-epoxy-1,2-butanediol, EBD) (Himmelstein et al., 1997; Melnick and
Kohn, 1995). Although there are quantitative differences in the
metabolic rates for various pathways between different species, the
metabolism of 1,3-butadiene is qualitatively similar among species. The
enzymes responsible for the metabolic activation of 1,3-butadiene to
these epoxide metabolites, as well as the enzymes responsible for the
detoxification of these reactive metabolites, exist in humans as well as
mice and rats. The genetic toxicology literature on 1,3-butadiene, EB,
and DEB consists of more than 450 publications with positive genotoxic
findings in viruses, bacteria, plants, and animals. EBD has been less
extensively studied, but recent evidence suggests that most of the
trihydroxybutyl-guanine adducts in mice and rats exposed to
1,3-butadiene are derived from EBD (Koc et al., 1999). In addition,
1,3-butadiene is structurally related to other (rodent) carcinogens,
such as isoprene and chloroprene (NTP, 1997; Melnick et al., 1996; NTP,
1998).

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

None. 1,3-Butadiene is a gas at room temperature and pressure, making
oral exposure unlikely.

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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. Inhalation Unit Risk - 3 × 10-5 per µg/m3 (0.08 per
ppm). 

____II.C.1.2. Extrapolation Method - linear extrapolation from LEC01
(0.254 ppm); LEC01 derived from linear relative rate model (RR = 1 + bX)
using lifetable analysis with leukemia incidence data; an adjustment
factor of 2 was applied (see below).

Air Concentrations at Specified Risk Levels:

Risk Level	Concentration

E-4 (1 in 10,000)	3 µg/m3

E-5 (1 in 100,000)	0.3 µg/m3

E-6 (1 in 1,000,000)	0.03 µg/m3

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

The Delzell et al. (1995) retrospective cohort study of more than 15,000
male styrene-butadiene rubber production workers provides high-quality
epidemiologic data on leukemia risk from 1,3-butadiene exposure. In the
Delzell et al. study, 1,3-butadiene exposure was estimated for each job
and work area for each study year, and these estimates were linked to
workers' work histories to derive cumulative exposure estimates for each
individual worker. Subsequent to the Poisson regression analyses by
Delzell et al., which used four different mathematical models (linear,
log-linear, power, and square root) to fit the exposure-response data,
Health Canada obtained the data on this cohort and performed their own
analyses. The Health Canada (1998) analyses were similar to those of
Delzell et al., but involved some minor refinements (e.g., finer
stratification of age and other modifying variables, and use of the
actual mean exposure in the highest exposure group rather than an
arbitrary value). It is the Health Canada analyses that are used for
this risk assessment. 

The linear relative rate model reported by Health Canada was RR = 1 +
0.0099X, where X represents cumulative 1,3-butadiene exposure in
ppm-years. The results were adjusted for age, calendar period, years
since hire, and cumulative styrene exposure. Benzene exposure was also
estimated for each worker but was not found to be a confounder, and
hence, was not included in the models. Risk estimates were made using
the relative rate models and an actuarial program that accounts for the
effects of competing causes of death. U.S. age-specific mortality rates
for all race and gender groups combined (NCHS, 1993) were used to
specify the leukemia and all-cause background rates. Risks were computed
up to age 85 for continuous 1,3-butadiene exposures. The occupational
exposures in the epidemiology study were converted to continuous
exposures by adjusting for the differences in the number of days exposed
per year (240/365 days) and differences in the amount of contaminated
air inhaled per day (10/20 m3). (10 m3 is the default occupational
ventilation volume for an 8-hour work shift; 20 m3 is the default
24-hour ambient ventilation volume [U.S. EPA, 1994]). 

Interpreting the proposed new carcinogen risk assessment guidelines
(U.S. EPA, 1999), linear extrapolation from the LEC01 (i.e., the 95%
lower confidence limit of the exposure concentration associated with a
1% increased risk) is warranted given the clear genotoxicity of
1,3-butadiene and the fact that a 1% increase in risk is within the
range of the epidemiologic data. Using the linear relative rate model
for modeling the epidemiologic data in the range of observation yields
an LEC01 of 0.375 ppm. Using the LEC01 as the point of departure and
extrapolating linearly to 0 increased risk at 0 exposure, a unit risk
estimate of 0.03/ppm is obtained for the risk of leukemia mortality from
the occupational data. 

However, we actually wish to estimate cancer incidence, not mortality;
therefore, another calculation was done using the linear relative rate
model and age-specific leukemia incidence rates for 1994-1998 from SEER
(Surveillance, Epidemiology and End Results program of the National
Cancer Institute; NCI, 2001) in place of the leukemia mortality rates in
the actuarial program. This calculation assumes that leukemia incidence
and mortality have the same exposure-response relationship for
1,3-butadiene exposure and that the incidence data are for first
occurrences of leukemia or that relapses provide a negligible
contribution. The calculation also relies upon the fact that the
leukemia incidence rates are small compared to the all-cause mortality
rates. The result is an LEC01 of 0.254 ppm and a unit risk estimate of
0.04/ppm for leukemia incidence. 

An adjustment factor of 2 was then applied to this unit risk estimate to
reflect evidence from rodent bioassays suggesting that extrapolating the
excess risk of leukemia in a male-only occupational cohort may
underestimate the total cancer risk from 1,3-butadiene exposure in the
general population. First, studies in both rats and mice indicate that
1,3-butadiene is a multisite carcinogen. It is possible that humans
exposed to 1,3-butadiene may also be at risk of cancers other than
leukemia and that the epidemiologic study had insufficient power to
detect excess risks at other sites (see below). Second, both the rat and
mouse studies suggest that females are more sensitive to
1,3-butadiene-induced carcinogenicity than males, and the female mammary
gland was the only 1,3-butadiene-related tumor site common to both
species. The mammary tumor unit risk estimated from the female mouse
(most sensitive species) data is just slightly lower (maximum likelihood
estimate [MLE] = 0.02/ppm, 95% upper confidence limit [UCL] = 0.03/ppm)
than the human (male) leukemia risk (0.04/ppm based on linear
extrapolation from the LEC01). Thus, EPA decided to apply an adjustment
factor of 2 to the leukemia risk estimate, resulting in a unit risk
estimate of 0.08/ppm intended to cover the combined risks for leukemia
and mammary cancer and to provide additional protection to account for
the fact that small increases in risk at other sites, particularly the
lung, cannot be ruled out. 

The Delzell et al. study was a large cohort study (over 15,000 subjects)
with a long follow-up time (49 years; 25% of the subjects had died by
the end of the follow-up), so for most tumor sites there should be
sufficient power to detect an increased risk. The main tumor site that
might be at issue is the lung, which was the most sensitive site in both
male and female mice. Lung cancer is fairly common in humans; therefore,
the epidemiology study may have lacked the power to detect an increase
in lung cancer. A crude "power" calculation based on the average
employment and exposure characteristics of the cohort, exposure
estimates and number of subjects available for 6 of the 8 plants, and
the MLE of lung cancer unit risk for the female mouse (i.e., 0.1/ppm),
and assuming no confounding by smoking, suggests that if humans were as
sensitive as mice to the lung cancer effects of 1,3-butadiene, one would
have expected to see 26 excess lung cancer cases in the epidemiology
study. In fact, only 2 excess lung cancer cases were observed in the
workers from the 6 plants over a background of 312 expected cases. On
the other hand, the study has low statistical power to detect such a
small proportional excess (power to detect a statistically significant
increase in risk if the true SMR = 338/312 = 108 is estimated to be 42%
according to the method of Beaumont and Breslow [1981]), and an SMR of
107 (319 observed/297 expected) was observed for the ever hourly workers
for the 8 plants (although there was no increased risk in the overall
cohort [SMR=101] or in the subgroup of ever hourly workers with >= 10
years worked and >= 20 years since hire [SMR=100]). 

The only process group associated with an increased lung cancer SMR was
maintenance (SMR = 141 observed/114 expected = 124). However, 7
mesotheliomas were also observed in maintenance workers (9 among ever
hourly workers in the total cohort), suggesting that these workers may
have been exposed to asbestos, a known lung carcinogen. Furthermore, the
evidence for the association between the increased lung cancers in the
maintenance workers and 1,3-butadiene exposure is weakened by the fact
that lung cancers are not increased in other process groups which
exhibited increases in leukemia cases (e.g., 1,3-butadiene production),
the absence of a positive relationship with number of years worked, the
absence of a trend with increasing years since hire, and the fact that
the increase was attenuated when state, rather than national, lung
cancer rates were used for comparison (Sathiakumar et al., 1998). Thus,
the overall evidence of an association between lung cancer and
1,3-butadiene exposure is tenuous. 

On the other hand, because the background rate of lung cancer is high,
the power of the study to detect small increases in lung cancer risk is
low, and, without adjusting for amount of smoking, it is difficult to
make firm conclusions. Workers are not allowed to smoke in the plants
because of the explosive potential of 1,3-butadiene; therefore, the
workers may have had lower cigarette consumption, and this could easily
mask a small increase in lung cancer risk. Thus, while the study does
not provide good evidence of an association between lung cancer and
1,3-butadiene exposure, one cannot rule out a small increase in risk. 

Some increases were also observed for laryngeal cancer in the Delzell et
al. study; however, these are based on small numbers (for the overall
cohort: 17 observed, 15 expected). On the other hand, the increases are
associated with process groups in which excess leukemias are observed.
No data are provided for duration of exposure or other exposure
characteristics. Thus, while the evidence for an association between
laryngeal cancer and 1,3-butadiene exposure is meager, a small increase
in laryngeal cancer cannot be ruled out.

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

For comparison purposes, human unit cancer risk estimates based on
extrapolation from the results of lifetime animal inhalation studies are
also presented. These unit risk estimates are 95% upper confidence
limits on unit cancer risk, calculated from incidence data on all
significantly elevated tumor sites using a linearized low-dose
extrapolation model, consistent with the 1986 guidelines (U.S. EPA,
1986). Exposure values were adjusted to 24-hour continuous equivalent
exposures by multiplying by (6/24) and (5/7). The rat-based estimates
are 4.2 × 10-3/ppm from male rat data and 5.6 × 10-2/ppm from female
rat data. These estimates are from EPA's 1985 assessment and were
derived using the linearized multistage model and estimates of absorbed
dose (U.S. EPA, 1985). The mouse-based estimates were derived from the
1993 NTP study, including interim kill data, using a Weibull multistage
time-to-tumor model and an assumption of ppm equivalence across species.
Unit risk estimates for each tumor type were calculated separately and a
Monte Carlo analysis was used to estimate the 95% upper bound on the sum
of the MLEs (U.S. EPA, 2002, Section 10.2.2.2). A cancer unit risk
estimate of 0.22/ppm was calculated from the male mouse data and
0.29/ppm from the female mouse data. The estimate of 0.3/ppm based on
the female mouse data is the preferred animal-based upper bound on human
risk. 

Human health risk estimates based on extrapolation from high-quality
epidemiologic results are preferable to those based on rodent data,
because they avoid the uncertainties inherent in extrapolating across
species and, typically, the human exposures in epidemiologic studies are
closer to anticipated environmental exposures than the high exposures
used in animal studies, thus reducing the extent of low-dose
extrapolation. In the case of 1,3-butadiene, while the rat exposures
were at least 100-fold higher than human exposures, the lowest exposure
in the 1993 NTP mouse study (4.7 ppm, 8-h TWA) is within the range of
occupational exposures (0.7-1.7 ppm median and 39-64 ppm max 8-hour TWAs
for work-area groups). However, interspecies differences in tumor sites
and susceptibilities between rats and mice are especially pronounced,
and the biological bases for these differences are unresolved. A review
of available pharmacokinetic data and models revealed that the state of
the science is currently inadequate for explaining interspecies
differences or improving on default dosimetry assumptions (see Section
I.B.4 and U.S. EPA, 2002, Chapter 9). Therefore, the quantitative
extrapolation of rodent risks to humans is highly uncertain for
1,3-butadiene.

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

Even though high-quality human data were used for the quantitative
cancer risk estimation for 1,3-butadiene, there are inevitable
uncertainties in the calculated risk estimate. First, there are
uncertainties inherent in the epidemiologic study itself. In particular,
there are uncertainties in the retrospective estimation of 1,3-butadiene
exposures, which could have resulted in exposure misclassification.
Nondifferential exposure misclassification would tend to bias estimates
of effect toward the null, resulting in an underestimate of risk.
Differential misclassification could bias results in either direction. 

In fact, after completing their initial study, Delzell et al. raised
some concerns about the accuracy of the exposure estimates (see U.S.
EPA, 2002, Section 10.1.3). In 2000, Delzell et al. completed a
re-assessment of the exposure estimates and concluded that the earlier a
priori estimates were too low. The revised exposure estimates need to be
critically evaluated before EPA can determine whether or not they are
more credible than the a priori estimates. If the revised exposure
estimates are valid, the leukemia portion of the cancer risk estimate
would decrease somewhat (see U.S. EPA, 2002, Section 11.1). 

Second, there are uncertainties regarding the appropriate dose metric
for dose-response analysis. Although the dose surrogate of cumulative
exposure (i.e., ppm × years) yielded highly statistically significant
exposure-response relationships, cumulative exposure is strongly
correlated with other possible exposure measures, and there may be a
dose-rate effect (e.g., risk at high exposures may be more than
proportionately greater than at lower exposures) obscured in the
analysis, or operative at exposures below the observable range but
relevant to low-dose extrapolation. 

Third, there are uncertainties pertaining to which model to use for the
epidemiologic data. Several mathematical models adequately fit the
exposure-response data from the epidemiology study, and because the
specific mechanisms of 1,3-butadiene carcinogenesis are unknown, there
is no biological basis for choosing one model over another. The linear
model was chosen in this risk assessment to derive the "point of
departure" for low-dose extrapolation because there was no compelling
reason to deviate from historical approaches. 

Fourth, it is uncertain which potential modifying or confounding factors
should be included in the model. The linear model of Health Canada,
which is used in this risk assessment, was adjusted for age, calendar
year, years since hire, race, and exposure to styrene. Plant and benzene
exposure were ruled out as potential confounders. However, there may be
other relevant factors that were not included in the models. 

Fifth, there are uncertainties in the parameter estimates used in the
models. The study of Delzell et al. is large, providing some degree of
reliability in the parameter estimates. However, especially given the
large human variability that has been observed in metabolic activities
that could affect cancer risk from 1,3-butadiene exposure, the
generalizability of the occupational results is unclear. 

Sixth, there are uncertainties in extending the relative rate models
from the epidemiology study to derive lifetime excess leukemia incidence
unit risk estimates for the U.S. population. Notwithstanding, the
actuarial-type analysis that was used is a well-established methodology,
and the background leukemia incidence rates and mortality rates used in
the analysis are from large national databases. 

Seventh, the precise model for low-dose extrapolation is unknown. The
linear default extrapolation procedure in the 1999 proposed guidelines
was used in this assessment because of the well-established genotoxicity
of 1,3-butadiene via its metabolites. 

In addition, there are important concerns raised by comparison with the
rodent data. First, the rodent studies suggest that 1,3-butadiene is a
multisite carcinogen. It is possible that humans may also be at risk of
1,3-butadiene-induced carcinogenicity at other sites and that the
epidemiologic study had insufficient power to detect the other excess
risks. In the mouse, for example, the lung is the most sensitive tumor
site. Significant excesses of lung cancer may not have been detectable
in the epidemiologic study because of the high background rates of lung
cancer in humans (see also II.C.2 above). Delzell et al. did observe a
slight increase in lung cancer among maintenance workers. The
epidemiology-based excess cancer risk estimate of 0.04/ppm, which is
based only on leukemias, may be an underestimate if other sites are also
at risk. 

Second, both the rat and mouse studies suggest that females are more
sensitive to 1,3-butadiene-induced carcinogenicity than are males, and
the mammary gland in females was the only tumor site common to both
species. If female humans are also more sensitive than males, then the
male-based risk estimates calculated from the epidemiology study would
underestimate risks to females. Because of these concerns, an adjustment
factor of 2 is used, as discussed above, yielding a cancer unit risk
estimate of 0.08/ppm. 

Despite these uncertainties, confidence in the excess cancer risk
estimate of 0.08/ppm is moderate. First, the estimate is based primarily
on human data. Furthermore, these data are from a large, high-quality
epidemiologic study in which 1,3-butadiene exposures were estimated for
each individual a priori to conducting the exposure-response analysis.
Although there are uncertainties in the exposure estimation, a serious
attempt was made to reconstruct historical exposures for specific tasks
and work areas at different time periods. It is virtually unprecedented
to have such a comprehensive exposure assessment for individual workers
in such a large occupational epidemiologic study. In addition, the
assumption of linearity for low-dose extrapolation is reasonable given
the clear evidence of genotoxicity by 1,3-butadiene metabolites. 

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

__II.D.1. EPA Documentation

Source Document — U.S. EPA, 2002 

This assessment was peer reviewed by external scientists. Their comments
have been carefully evaluated and incorporated in finalization of this
IRIS Summary. A record of these comments is included as an appendix to
U.S. EPA, 2002.

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

Agency Consensus Date — 9/13/2001

__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  (email address).

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_III.  [reserved]

_IV.  [reserved] 

_V.  [reserved]

_VI.  Bibliography 

Substance Name — 1,3-Butadiene

CASRN — 106-99-0

Last Revised — 11/05/2002

_VI.A. Oral RfD References

None

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_VI.B. Inhalation RfC References

Adler, ID; Anderson, D. (1994) Dominant lethal effects after inhalation
exposure to 1,3-butadiene. Mutat Res 309:295-297. 

Anderson, D; Edwards, AJ; Brinkworth, MH. (1993) Male-mediated F1
effects in mice exposed to 1,3-butadiene. In: Butadiene and styrene:
assessment of health hazards. IARC Scientific Publications. Vol. 127.
Sorsa, M; Peltonen, K; Vainio, H; et al., eds. Lyon, France:
International Agency for Research on Cancer, pp. 171-181. 

Anderson, D; Hughes, JA; Edwards, AJ; et al. (1998) A comparison of
male-mediated effects in rats and mice exposed to 1,3-butadiene. Mutat
Res 397:77-84. 

Bevan, C; Stadler, JC; Elliot, GS; et al. (1996) Subchronic toxicity of
4-vinylcyclohexene in rats and mice by inhalation. Fundam Appl Toxicol
32:1-10. 

Brinkworth, MH; Anderson, D; Hughes, JA; et al. (1998) Genetic effects
of 1,3-butadiene on the mouse testis. Mutat Res 397:67-75. 

Hackett, PL; Sikov, MR; Mast, TJ; et al. (1987) Inhalation developmental
toxicology studies: teratology study of 1,3-butadiene in mice (final
report). Richland, WA: Pacific Northwest Laboratory; PNL Report No.
PNL-6412 UC-48; NIH Report No. NIH-401-ES-40131; 92. Prepared for NIEHS,
NTP, under a Related Services Agreement with the U.S. Department of
Energy under contract DE-AC06-76RLO-1830. 

National Center for Health Statistics (NCHS). (1996) Vital statistics of
the United States, 1992, vol. II, mortality, part A. Washington, DC:
U.S. Public Health Service. 

National Toxicology Program (NTP), U.S. Department of Health and Human
Services. (1993) Toxicology and carcinogenesis studies of 1,3-butadiene
(CAS No. 106-99-0) in B6C3F1 mice (inhalation studies). NTP TR 434, NIH
Pub. No. 93-3165. Research Triangle Park, NC. 

U.S. Environmental Protection Agency (U.S. EPA). (1985) Mutagenicity and
carcinogenicity assessment of 1,3-butadiene. Office of Health and
Environmental Assessment, Office of Research and Development,
Washington, DC. EPA/600/8-85/004F. 

U.S. Environmental Protection Agency (U.S. EPA). (1994) Methods for
Derivation of Inhalation Reference Concentrations and Application of
Inhalation Dosimetry. Office of Research and Development, Washington,
DC. EPA/600/8-90/066F. 

U.S. Environmental Protection Agency (U.S. EPA). (2002) Health
assessment document for 1,3-butadiene. Office of Research and
Development, Washington, DC. EPA/600/P-98/001.

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_VI.C. Carcinogenicity Assessment References

Beaumont, JJ; Breslow, NE. (1981) Power considerations in epidemiologic
studies of vinyl chloride workers. Am J Epidemiol 114:752-34. 

Berard, CW; Hutchison, RE. (1997) The problem of classifying lymphomas:
an orderly prescription for progress. Ann Oncol (Suppl 2):53-59. 

Cowles, SR; Tsai, SP; Snyder, PJ; et al. (1994) Mortality, morbidity,
and haematologic results from a cohort of long-term workers involved in
1,3-butadiene monomer production. Occup Environ Med 51:323-329. 

Delzell, E; Sathiakumar, N; Macaluso, M.; et al. (1995) A follow-up
study of synthetic rubber workers. Submitted to the International
Institute of Synthetic Rubber Producers. University of Alabama at
Birmingham. October 2, 1995. 

Delzell, E; Sathiakumar, N; Hovinga, M. (1996) A follow-up study of
synthetic rubber workers. Toxicology 113:182-189. 

Divine, BJ; Hartman, CM. (1996) Mortality update of butadiene production
workers. Toxicology 113:169-181. 

Downs, TD; Crane, MM; Kim, KW. (1987) Mortality among workers at a
butadiene facility. Am J Ind Med 12:311-329. 

Hazleton Laboratories Europe Ltd. (1981) The toxicity and
carcinogenicity of butadiene gas administered to rats by inhalation for
approximately 24 months. Prepared for the International Institute of
Synthetic Rubber Producers, New York, NY. Unpublished. 

Health Canada. (1998) Canadian Environmental Protection Act Priority
Substances List Health Assessment: 1,3-Butadiene. Draft for second stage
peer review. Ottawa. March 1998. Additional information was provided by
Health Canada for calculation of the confidence intervals (personal
communication from Michael Walker to Leslie Stayner, 18 June 1999). 

Himmelstein, MW; Acquavella, JF; Recio, L; et al. (1997) Toxicology and
epidemiology of 1,3-butadiene. Crit Rev Toxicol 27(1):1-108. 

Irons, RD; Cathro, HP; Stillman WS; et al. (1989) Susceptibility to
1,3-butadiene-induced leukemogenesis correlates with endogenous
ecotropic retroviral background in the mouse. Toxicol Appl Pharmacol
101:170-176. 

Irons, RD; Pyatt, DW. (1998) Commentary: dithiocarbamates as potential
confounders in butadiene epidemiology. Carcinogenesis 19(4):539-542. 

Koc, H; Tretyakova, NY; Walker, VE; et al. (1999) Molecular dosimetry of
N-7 guanine adduct formation in mice and rats exposed to 1,3-butadiene.
Chem Res Toxicol 12:566-574. 

Macaluso, M; Larson, R; Delzell, E. (1996) Leukemia and cumulative
exposure to butadiene, styrene and benzene among workers in the
synthetic rubber industry. Toxicology 113:190-202. 

Matanoski, GM; Schwartz, L. (1987) Mortality of workers in
styrene-butadiene polymer production. J Occup Med 29:675-680. 

Melnick, RL; Kohn, MC. (1995) Mechanistic data indicate that
1,3-butadiene is a human carcinogen. Carcinogenesis 16(2):157-163. 

Melnick, RL; Elwell, MR; Roycroft, JH; et al. (1996) Toxicity of inhaled
chloroprene (2-chloro-1,3-butadiene) in F344 rats and B6C3F1 mice.
Toxicology 113:247-252. 

National Cancer Institute (NCI). (2001) SEER Cancer Statistics Review,
1973-1998, National Cancer Institute. Bethseda, MD,   HYPERLINK
"http://seer.cancer.gov/csr/1973_1998/index.html" 
http://seer.cancer.gov/csr/1973_1998/index.html , 2001. 

National Center for Health Statistics (NCHS). (1996) Vital statistics of
the United States, 1992, vol. II, mortality, part A. Washington, DC:
U.S. Public Health Service. 

National Toxicology Program (NTP), U.S. Department of Health and Human
Services. (1984) Toxicology and carcinogenesis studies of 1,3-butadiene
(CAS No. 106-99-0) in B6C3F1 mice (inhalation studies). NTP TR 288, NIH
Pub. No. 84-2544. Research Triangle Park, NC. 

NTP. (1993) Toxicology and carcinogenesis studies of 1,3-butadiene (CAS
No. 106-99-0) in B6C3F1 mice (inhalation studies). NTP TR 434, NIH Pub.
No. 93-3165. Research Triangle Park, NC. 

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U.S. DHHS. 

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F344/N rats and B6C3F1 mice. Technical Report No. 467, U.S. DHHS. 

Owen PE; Glaister, JR; Gaunt, IF; et al. (1987) Inhalation toxicity
studies with 1,3-butadiene. 3. Two-year toxicity/carcinogenicity study
in rats. Am Ind Hyg Assoc J 48(5):407-413. 

Santos-Burgoa, C; Matanoski, GM; Zeger, S; et al. (1992)
Lymphohematopoietic cancer in styrene-butadiene polymerization workers.
Am J Epidemiol 136:843-854. 

Sathiakumar, N; Delzell, E; Hovinga, M; et al. (1998) Mortality from
cancer and other causes of death among synthetic rubber workers. Occup
Environ Med 55:230-235. 

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Washington, DC. EPA/600/8-85/004F. 

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"http://www.epa.gov/ncea/raf/cancer.htm" 
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cohort. Environ Health Perspect 103:598-603.

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_VII.  Revision History

Substance Name — 1,3-Butadiene

CASRN — 106-99-0

Date	Section	Description

03/01/1988	II.A.2.	Text revised

03/01/1988	II.C.4.	Confidence statement revised

06/01/1989	II.D.3.	Primary contact changed

07/01/1989	II.A.2.	Correct Meinhardt citation

07/01/1989	II.A.4.	Correct citations

07/01/1989	VI.	Bibliography on-line

06/01/1990	IV.A.1.	Area code for EPA contact corrected

01/01/1991	II.	Text edited

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

01/01/1991	VI.C.	Bond et al., 1986 and Cote and Bayard, 1990 refs added

02/01/1991	II.C.3.	Information on extrapolation process included

01/01/1992	IV.	Regulatory actions updated

04/01/1992	VI.	Regulatory action section withdrawn

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.

01/02/1998	I.,II.	This chemical is being reassessed under the IRIS
Program.

11/05/2002	All	Complete revision based on new health assessment document

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_VIII.  Synonyms

Substance Name — 1,3-Butadiene

CASRN — 106-99-0

Last Revised — 11/05/2002

106-99-0 

BIETHYLENE 

BIVINYL 

BUTADIEEN 

BUTA-1,3-DIEEN 

BUTADIEN 

BUTA-1,3-DIEN 

BUTADIENE 

1,3-Butadiene 

Butadiene, 1,3- 

alpha,gamma-BUTADIENE 

DIVINYL 

ERYTHRENE 

NCI-C50602 

PYRROLYLENE 

VINYLETHYLENE