United States
kS^laMIjk Environmental Protection
^J^iniiil m11 Agency
EPA/690/R-10/002F
Final
9-13-2010
Provisional Peer-Reviewed Toxicity Values for
ft-Butylbenzene
(CASRN 104-51-8)
Superfund Health Risk Technical Support Center
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, OH 45268

-------
TABLE OF CONTENTS
COMMONLY USED ABBREVIATIONS	ii
BACKGROUND	1
HISTORY	1
DISCLAIMERS	1
QUESTIONS REGARDING PPRTVS	2
INTRODUCTION	2
REVIEW 01 PERTINENT DATA	2
HUMAN STUDIES	2
ANIMAL STUDIES	3
OTHER STUDIES	3
DERIVATION OF PROVISIONAL SI BCI IRONIC AND CHRONIC	4
ORAL RID VALUES I OR w-BUTYLBENZENE	4
SI BCI IRONIC p-RlD	4
CHRONIC p-RlD	6
FEASIBILITY OF DERIVING PROVISIONAL SUBCHRONIC AND CHRONIC INHALATION RfC
VALUES FOR w-BUTYLBENZENE	7
PROVISIONAL CARCINOGENICITY ASSESSMENT	7
FOR w-BUTYLBENZENE	7
WEIGHT-OF -E VIDEN CE DESCRIPTOR	7
QUANTITATIVE ESTIMATES OF CARCINOGENIC RISK	7
REFERENCES	7
APPENDIX A. DETAILS OF BENCHMARK DOSE MODELING FOR SUBCHRONIC AND
CHRONIC p-RlDs	10
l

-------
COMMONLY USED ABBREVIATIONS
BMC
benchmark concentration
BMD
benchmark dose
BMCL
benchmark concentration lower bound 95% confidence interval
BMDL
benchmark dose lower bound 95% confidence interval
HEC
human equivalent concentration
HED
human equivalent dose
IUR
inhalation unit risk
LOAEL
lowest-observed-adverse-effect level
LOAELadj
LOAEL adjusted to continuous exposure duration
LOAELhec
LOAEL adjusted for dosimetric differences across species to a human
NOAEL
no-ob served-adverse-effect level
NOAELadj
NOAEL adjusted to continuous exposure duration
NOAELhec
NOAEL adjusted for dosimetric differences across species to a human
NOEL
no-ob served-effect level
OSF
oral slope factor
p-IUR
provisional inhalation unit risk
p-OSF
provisional oral slope factor
p-RfC
provisional reference concentration (inhalation)
p-RfD
provisional reference dose (oral)
POD
point of departure
RfC
reference concentration (inhalation)
RfD
reference dose (oral)
UF
uncertainty factor
UFa
animal-to-human uncertainty factor
UFC
composite uncertainty factor
UFd
incomplete-to-complete database uncertainty factor
UFh
interhuman uncertainty factor
UFl
LOAEL-to-NOAEL uncertainty factor
UFS
subchronic-to-chronic uncertainty factor
WOE
weight of evidence
11

-------
FINAL
9-13-2010
PROVISIONAL PEER-REVIEWED TOXICITY VALUES FOR
ft-BUTYLBENZENE (CASRN 104-51-8)
BACKGROUND
HISTORY
On December 5, 2003, the U.S. Environmental Protection Agency's (EPA) Office of
Superfund Remediation and Technology Innovation (OSRTI) revised its hierarchy of human
health toxicity values for Superfund risk assessments, establishing the following three tiers as the
new hierarchy:
1)	EPA's Integrated Risk Information System (IRIS).
2)	Provisional Peer-Reviewed Toxicity Values (PPRTVs) used in EPA's Superfund
Program.
3)	Other (peer-reviewed) toxicity values, including
~	Minimal Risk Levels produced by the Agency for Toxic Substances and Disease
Registry (ATSDR),
~	California Environmental Protection Agency (CalEPA) values, and
~	EPA Health Effects Assessment Summary Table (HEAST) values.
A PPRTV is defined as a toxicity value derived for use in the Superfund Program when
such a value is not available in EPA's IRIS. PPRTVs are developed according to a Standard
Operating Procedure (SOP) and are derived after a review of the relevant scientific literature
using the same methods, sources of data, and Agency guidance for value derivation generally
used by the EPA IRIS Program. All provisional toxicity values receive internal review by two
EPA scientists and external peer review by three independently selected scientific experts.
PPRTVs differ from IRIS values in that PPRTVs do not receive the multiprogram consensus
review provided for IRIS values. This is because IRIS values are generally intended to be used
in all EPA programs, while PPRTVs are developed specifically for the Superfund Program.
Because new information becomes available and scientific methods improve over time,
PPRTVs are reviewed on a 5-year basis and updated into the active database. Once an IRIS
value for a specific chemical becomes available for Agency review, the analogous PPRTV for
that same chemical is retired. It should also be noted that some PPRTV documents conclude that
a PPRTV cannot be derived based on inadequate data.
DISCLAIMERS
Users of this document should first check to see if any IRIS values exist for the chemical
of concern before proceeding to use a PPRTV. If no IRIS value is available, staff in the regional
Superfund and Resource Conservation and Recovery Act (RCRA) program offices are advised to
carefully review the information provided in this document to ensure that the PPRTVs used are
appropriate for the types of exposures and circumstances at the Superfund site or RCRA facility
in question. PPRTVs are periodically updated; therefore, users should ensure that the values
contained in the PPRTV are current at the time of use.
1

-------
FINAL
9-13-2010
It is important to remember that a provisional value alone tells very little about the
adverse effects of a chemical or the quality of evidence on which the value is based. Therefore,
users are strongly encouraged to read the entire PPRTV document and understand the strengths
and limitations of the derived provisional values. PPRTVs are developed by the EPA Office of
Research and Development's National Center for Environmental Assessment, Superfund Health
Risk Technical Support Center for OSRTI. Other EPA programs or external parties who may
choose of their own initiative to use these PPRTVs are advised that Superfund resources will not
generally be used to respond to challenges of PPRTVs used in a context outside of the Superfund
Program.
QUESTIONS REGARDING PPRTVS
Questions regarding the contents of the PPRTVs and their appropriate use (e.g., on
chemicals not covered, or whether chemicals have pending IRIS toxicity values) may be directed
to the EPA Office of Research and Development's National Center for Environmental
Assessment, Superfund Health Risk Technical Support Center (513-569-7300), or OSRTI.
INTRODUCTION
No RfD, RfC, or carcinogenicity assessment for //-butylbenzene is available on IRIS
(U.S. EPA, 2008), in the Drinking Water Standards and Health Advisories list (U.S. EPA, 2006),
or in the HEAST (U.S. EPA, 1997). The only document on the CARA list (U.S. EPA, 1991,
1994) that includes information about //-butylbenzene is a Drinking Water Health Advisory for
//-Butylbenzene (U.S. EPA, 1987); it concluded that data were inadequate for derivation of
health advisory levels. ATSDR (2008) has not produced a Toxicological Profile for
^-butylbenzene, and no Environmental Health Criteria Document is available (WHO, 2008).
The American Conference of Governmental Industrial Hygienists (ACGIH, 2007), the
Occupational Safety and Health Administration (OSHA, 2008), and the National Institute for
Occupational Safety and Health (NIOSH, 2008) have not established occupational health
standards for ^-butylbenzene. The carcinogenicity of ^-butylbenzene has not been assessed by
IARC (2008) or NTP (2005, 2008).
Literature searches were conducted from the 1960s through December 2009 for studies
relevant to the derivation of provisional toxicity values for //-butylbenzene. Databases searched
include MEDLINE, TOXLINE (Special), BIOSIS, TSCATS1/TSCATS 2, CCRIS, DART/ETIC,
GENETOX, HSDB, RTECS, and Current Contents. A review by Henderson (2001) was also
consulted for relevant information.
REVIEW OF PERTINENT DATA
HUMAN STUDIES
No relevant human studies were located.
2

-------
FINAL
9-13-2010
ANIMAL STUDIES
In a two-generation reproductive summary report (Yamasaki et al., 2005) in SD rats for
nine chemicals, ^-butylbenzene was administered orally by gavage at dose levels of 0, 30, 100,
or 300 mg/kg-day. The study authors found no effects on the endocrine system and no
reproductive effects in the F0 and F1 parents or F1 and F2 offspring for the study of
//-butylbenzene. The study was primarily designed for detecting endocrine-mediated influence
by //-butylbenzene and other eight chemicals; the study authors conducted no further
histopathology or clinical chemistry on the rats exposed to ^-butylbenzene at any dose level.
The highest tested dose in this study, 300 mg/kg-day, is considered a NOAEL in this PPRTV
document.
In a preliminary study, ^-butylbenzene (98% purity) was administered by gavage to
Crj:CD (SD) IGS rats (6/sex/dose) at doses of 0, 30, 100, 300, or 1000 mg/kg-day every day for
4 weeks prior to mating, and throughout gestation and lactation (Izumi et al., 2005). Methods for
the preliminary study are not further described. Body-weight gain was inhibited in parental rats
treated with 1000 mg/kg-day, and the study authors evaluated the liver, kidney, and adrenal
weights in males treated with 300 or 1000 mg/kg-day (no further details reported; data not
shown). The study authors reported that there were no effects on the fertility of the parental
animals, but there was a decreased viability of offspring at Postnatal Day (PND) 4 among F1 rats
in the 1000-mg/kg-day group; body-weight gain was also inhibited among F1 offspring in this
treatment group (no further details reported; data not shown).
Based on the results of this study, ^-butylbenzene was administered by gavage (in olive
oil; volume adjusted to 5 ml/kg) to Crj:CD (SD) IGS rats (24/sex/dose) at doses of 0, 30, 100, or
300 mg/kg-day, every day for over two generations (Izumi et al., 2005). F0 males and females
were exposed for 10 weeks prior to mating and during the mating period (up to 2 weeks).
F0 males were exposed for an additional 4-6 weeks after mating, while F0 females were exposed
throughout gestation and lactation (to Day 21); both sexes were exposed for a total of
14-16 weeks. F1 males were exposed for approximately 18 weeks, and F1 females were
exposed for 19-21 weeks; exposures included the 10 weeks prior to mating (starting at weaning
at 3 weeks of age), the mating period, and, in females, gestation and lactation (to Day 21).
Animals were observed daily for clinical signs and mortality. Body weight and food
consumption were measured weekly for males, and weekly for females prior to conception. For
females, body weight was also measured on Gestational Days (GDs) 0, 7, 14, and 20, and PNDs
0, 4, 7, 14, and 21. Food consumption was measured on GDs 1,7, 14 and 20, and on PNDs 1, 4,
7, 14, and 21. Fertility was assessed in parental animals by measurement of estrous count,
estrous interval, number of pregnancies, number of confirmed copulations, number of viable
offspring, implantations, gestation length, and litter size. The timing of sexual maturation was
assessed in offspring by examining them for vaginal opening (females) or preputial separation
(males) every day until completion of the process. Sperm counts and motility were determined
for 10 males per dose. Hormone levels, including estradiol, testosterone, follicle stimulating
hormone (FSH), and luteinizing hormone (LH) were assessed in six males (different group than
those in whom sperm variables were assessed) and six females from each dose.
The following variables were assessed for each litter: number of each sex, number of live
offspring, and number of live offspring on PNDs 4 and 21 (Izumi et al., 2005). Offspring were
assessed daily for clinical signs. Each litter was culled to four males and four females, where
3

-------
FINAL
9-13-2010
possible, on PND 4. The following developmental milestones were assessed for each individual
within a litter to detect potential effects related to endocrine disruption: pinna detachment from
PND 4 to completion; incisor eruption from PND 10 to completion; eyelid separation from PND
15 to completion; righting reflex from PND 5 to completion; visual placing reflex from PND 16
to completion and Preyer's reflex from PND 28 to completion. Organ weights were assessed for
all major organs in parental animals. The weights of brain, thymus, spleen, testes, epididymides,
ovaries, and uterus were measured in F1 and F2 weanlings. All parental animals, offspring that
died during lactation or were culled and weanlings that were not selected as parental animals
were necropsied. The following tissues from all F0 and F1 adult animals from the control and
high-dose groups were examined microscopically: pituitary gland, thyroid and parathyroid
glands, liver, kidneys, adrenal glands, testes, epididymides, seminal vesicles, coagulating glands,
prostate, ovaries, oviduct, uterus, uterine cervix, vagina, mammary glands, and any
macroscopically identified abnormal tissue. Based on changes observed in the high-dose group,
liver, kidneys, ovaries, oviduct, uterus, uterine cervix, and vagina were also examined
microscopically for the 30- and 100-mg/kg-day dose groups. The pituitary glands, testes,
epididymides, prostate, seminal vesicles, and coagulating glands were examined in the
noncopulating and infertile copulating males in the 30-and 100-mg/kg-day dose groups and in
five fertile animals from the control group.
No treatment-related mortality was observed in parental animals, although one high-dose
F0 female and one control F1 female died during the study due to spontaneous leukemia and
gavage error, respectively (Izumi et al., 2005). Excessive salivation was observed immediately
after //-butylbenzene administration in parental rats (primarily the males) from both generations
in the 100- and 300-mg/kg-day treatment groups. No other clinical signs were observed.
According to the study authors, body weight and food consumption were not affected by
treatment, and there were no treatment-related effects on any indicator of fertility1 in either
parental sex and in any generation (but, in the F0 dams, there was a tendency for reduced number
of implantations at 300 mg/kg/day and a prolonged estrus interval at 100 and 300 mg/kg/day but
not statistically significant \p < 0.05]), and no treatment-related effects on hormone levels of
parental animals. No treatment-related effects were observed upon gross necropsy of parental
animals. Statistically significant increases (12-19% at/? < 0.05) in absolute and relative liver
weight were seen in parental F0 males at >100 mg/kg-day (liver histopathological changes were
observed at 300 mg/kg-day, but not at 100 mg/kg-day), F0 females at >30 mg/kg-day, and
F1 males and females at 300 mg/kg-day (no histopathological change was observed in the liver
of the female animals). Absolute and relative kidney weights were statistically significantly
increased (7-21% at/? < 0.05) in parental male (with histopathological changes) and female rats
(with no histopathological change) of both generations at 300 mg/kg-day. In addition, a
histopathological change accompanied by statistically increased relative kidney weights (7% at
p < 0.05) at 100 mg/kg-day was observed in F1 parent males only. As discussed by Izumi et al.
(2005), the observed pattern of kidney changes is consistent with male-rat-specific
alpha2U-globulin-associated nephropathy—a condition that is not relevant to humans, but an
analysis of the mode of action has not been conducted in this assessment. Therefore, this effect
is considered relevant to humans. Incidence data for liver and kidney histopathological effects
are presented in Table 1. There was also a statistically significant increase (13% at p < 0.05) in
absolute and relative adrenal glands weights in high-dose F1 parental females (but not F1 males
1 Estrous count, estrous interval, copulations, numbers mating, number of pregnant females, gestation length,
numbers of implantations, litter size, etc.
4

-------
FINAL
9-13-2010
or F0 males or females). There were no histopathological findings in adrenal glands of parental
animals of either generation. There were no other significant treatment-related effects on organ
weights, including reproductive organs, in F0 or F1 parental animals.
Table 1. Incidence of Histopathological Findings of Interest in Parental Male Rats via Oral
Exposure to tt-Butylbenzene



Mean Dose (mg/kg-day)
Generation/Sex/T arget/Lesion
0
30
100
300
F0 Males
Liver, Hypertrophy, hepatocytes
0/24
0/24
0/24
5/24a
Kidney, hyaline droplets, proximal tubules
0/24
0/24
1/24
1 l/24b
Kidney, basophilic tubules
0/24
0/24
0/24
5/24a
F1 Males
Liver, Hypertrophy, hepatocytes
0/19
0/19
0/21
6/19b
Kidney, hyaline droplets, proximal tubules
0/19
1/19
5/2 r
12/193
Kidney, basophilic tubules
0/19
0/19
1/21
5/19a
aSignificantly different from controls (p < 0.05).
bSignificantly different from controls (p < 0.01).
Source: Izumi et al. (2005).
In summary, for parental animals several effects were observed (increased liver, kidney,
and adrenal weights, hyaline droplets in proximal tubules, and hepatocellular hypertrophy) at
300 mg/kg-day. Statistically significant increases in liver weight were seen in parental F0 males
at >100 mg/kg-day (liver histopathological changes were observed at 300 mg/kg-day, but not at
100 mg/kg-day), F0 females at >30 mg/kg-day, and F1 males and females at 300 mg/kg-day (no
histopathological change was observed in the liver of the female animals). Absolute and relative
kidney weights were statistically significantly increased in parental male (with histopathological
changes) and female rats (with no histopathological change) of both F0 and F1 generations at
300 mg/kg-day. A histopathological change accompanied by statistically increased relative
kidney weights (7%) at 100 mg/kg-day was observed in F1 parent males only. There was also a
statistically significant increase in absolute and relative adrenal glands weights in F1 parental
females (but not F1 males or F0 males or females) at 300 mg/kg-day that was not accompanied
by histopathological changes. The LOAEL for parental animals is identified as 300 mg/kg-day
based on hepatocellular hypertrophy, and increases in liver, kidney, and adrenal weights.
There were no treatment-related effects on any measure of growth or development in
F1 or F2 offspring, other than slight increases in absolute and/or relative thymus weight in some
groups exposed to 300 mg/kg-day (Izumi et al., 2005). Thymus and body-weight data are shown
in Table 2. The magnitude of the observed increases ranged from 10-27%, but the standard
deviations around the control means were large (15-32%), so, that in each case, the observed
increase was within 1 standard deviation of the control mean. There were no consistent changes
in other organ weights in the F1 or F2 pups (liver and kidney weights not reported).
5

-------
FINAL
9-13-2010
Table 2. Body and Thymus Weights in Rats
Treated with n-Butylbenzene
Parameter
Dose (mg/kg-day)
0
30
100
300
Males
Females
Males
Females
Males
Females
Males
Females
F0 parents
Number of rats
24
19
24
19
24
21
24
19
Body weight (g)
636.7 ±72.9a
322.8 ±24.7
627.6 ±51.2
331.3 ±23.7
635.09 ±70.2
328.4 ± 17.9
623.4 ±60.8
321.4 ±22.6
Liver weight
Absolute (g)
Relative (%)
21.35 ±3.71
3.34 ±0.28
13.30 ± 1.62
4.13 ±0.47
21.53 ±2.54
3.43 ±0.21
15.09 ± 1.50b
4.56 ± 0.33°
22.78 ±3.04
3.58 ±0.18
15.11 ± 1.64b
4.61 ± 0.50b
25.02 ±3.42b
4.00 ± 0.27b
15.20 ± 1.70b
4.74 ± 0.54b
Thymus weight
Absolute (g)
Relative (%)
0.27 ±0.08
0.04 ±0.01
0.21 ±0.06
0.07 ± 0.02
0.26 ±0.08
0.04 ±0.01
0.22 ±0.05
0.06 ±0.02
0.27 ±0.09
0.04 ±0.01
0.21 ±0.07
0.06 ± 0.02
0.25 ±0.07
0.04 ±0.01
0.22 ± 0.07
0.07 ± 0.02
F1 parents
Number of rats
19
12
19
11
21
12
19
13
Body weight (g)
680.4 ±99.7
328.9 ±21.9
698.4 ±45.1
329.2 ±20.7
668.0 ±61.3
323.4 ±33.6
653.3 ±75.5
325.4 ±20.3
Liver weight
Absolute (g)
Relative (%)
24.05 ±5.57
3.51 ±0.34
15.15 ± 1.75
4.62 ±0.52
25.08 ±2.95
3.58 ±0.25
15.03 ± 1.44
4.57 ±0.37
24.15 ±2.75
3.62 ±0.24
14.90 ± 1.52
4.62 ±0.39
26.82 ±4.14
4.09 ± 0.28b
17.02 ± 2.26°
5.22 ± 0.48b
Thymus weight
Absolute (g)
Relative (%)
0.29 ±0.08
0.04 ±0.01
0.22 ± 0.07
0.07 ± 0.02
0.29 ±0.07
0.04 ±0.01
0.22 ±0.07
0.07 ±0.02
0.30 ±0.08
0.04 ±0.01
0.22 ± 0.07
0.07 ± 0.02
0.28 ±0.12
0.04 ± 0.02
0.21 ±0.06
0.06 ± 0.02
F1 offspring
Number of rats
19
19
19
19
21
21
17
17
Body weight (g)
61.4 ±6.6
57.6 ± 10.8
61.5 ±6.0
59.9 ±5.0
59.1 ± 6.1
57.3 ±5.9
65.3 ±7.6
64.0 ±6.0
Thymus weight
Absolute (g)
Relative (%)
0.22 ± 0.07
0.36 ±0.07
0.23 ± 0.07
0.40 ±0.08
0.24 ± 0.04
0.39 ±0.05
0.24 ±0.04
0.40 ±0.06
0.21 ±0.04
0.35 ±0.05
0.21 ±0.05
0.37 ±0.07
0.28 ± 0.04°
0.43 ± 0.05b
0.28 ± 0.06°
0.44 ±0.08
1

-------
FINAL
9-13-2010
Table 2. Body and Thymus Weights in Rats
Treated with n-Butylbenzene
Parameter
Dose (mg/kg-day)
0
30
100
300
Males
Females
Males
Females
Males
Females
Males
Females
F2 offspring
Number of rats
12
12
11
11
12
12
13
13
Body weight (g)
71.4 ±6.2
68.3 ±5.8
71.9 ± 3.3
66.1 ±5.5
66.6 ±5.5
64.9 ±7.1
71.8 ±4.1
67.2 ±6.3
Thymus weight
Absolute (g)
Relative (%)
0.29 ±0.06
0.40 ± 0.06
0.25 ± 0.05
0.37 ±0.07
0.30 ±0.05
0.41 ±0.07
0.28 ±0.05
0.41 ±0.07
0.26 ± 0.06
0.38 ±0.08
0.26 ± 0.06
0.40 ± 0.06
0.29 ±0.03
0.41 ±0.05
0.29 ±0.03
0.43 ± 0.05°
aMean ± standard deviation.
bSignificantly different from controls (p < 0.01).
Significantly different from controls (p < 0.05).
N/A = not applicable.
Source: Izumi et al. (2005). Liver weight is not reported for F1/F2 offspring.
2

-------
FINAL
9-13-2010
The authors considered the effects on pup thymus weight to be treatment-related, but there were
some inconsistencies that suggest that the observed changes may not be toxicologically relevant. The
only evidence clearly supporting an effect is the statistically significant increase (p < 0.05) in both
absolute and relative thymus weight in F1 male pups at the high-dose of 300 mg/kg-day. Statistically
significant increases (p < 0.05) in absolute—but not relative—thymus weight in F1 female pups at
300 mg/kg-day, and relative—but not absolute—thymus weight in F2 female pups at 300 mg/kg-day
offer only ambiguous support. These changes were not internally consistent (i.e., absolute and relative
weights were not both changed together) and were proportional to body weight changes observed in the
same groups (nonsignificant increase in body weight in F1 female pups and nonsignificant decrease in
F2 female pups).
Overall, the biological significance of the change in pup thymus weight is questionable, but if
one assumes that the effect is a significant effect, then there is a potential concern for immunotoxicity
and no immunological assays have been conducted. Increased thymus weight (10-27%) in young
offspring may be an indicator of immunotoxicity because the thymus gland is a key organ for the
immune system (i.e., processing and maturation of T-cells). Despite the lack of histological findings in
parental animals (including the F1 parents that received in utero exposure), and the small increases in
mean organ weights observed in the weanlings, the observed effect on thymus weight in the weanlings is
considered to be biologically significant for this assessment in the absence of data to indicate otherwise.
A LOAEL of 300 mg/kg-day based on the increased thymus weight in F2 females is identified. The
NOAEL is 100 mg/kg-day.
Izumi et al. (2005) concluded that hepatocellular hypertrophy and increases in liver weight in
parental rats are an adaptive, rather than adverse, effect of //-butylbenzene on the liver based on
enzymatic induction of rat liver cytochrome P450 at an equivalent dose of 670 mg/kg as demonstrated
by Imaoka and Funane (1991). The effect of hepatocellular hypertrophy may be specific to males
because no histological change was observed in the liver of the female rats (but the liver weight in
females were increased significantly at lower doses). Overall, these liver effects cannot be discounted
because there is uncertainty of the enzymatic induction at the low-dose region (<300 mg/kg-day) and its
potential extrapolation and relevance to humans. For this review, these liver effects are considered
biologically significant, and a LOAEL is established at 300 mg/kg-day based on the hepatocellular
hypertrophy and is supported by (absolute and/or relative) increased liver weight in F0/F1 parent males.
The NOAEL is 100 mg/kg-day.
OTHER STUDIES
Tanii et al. (1995) reported an i.p. LD50 of 1.995 g/kg for ^-butylbenzene in mice. Following
acute oral exposure to 4.3 g/kg //-butylbenzene, 2/10 rats died (Gerarde, 1959). Lethality was higher for
branched-chain butylbenzenes in this study (8/10 died for sec-butylbenzene and 7/10 for
/^/'/-butylbenzene at the same dose). The leading cause of death in rats in this study was chemical-
induced pneumonitis with pulmonary edema and hemorrhage, the latter often associated with
hemorrhage in other tissues such as thymus, adrenal, and bladder. The study authors also reported
hyperemia and vasodilation of the blood vessels of the gastrointestinal tract.
Noting that aromatic solvents including toluene, ethylbenzene, styrene, and /^-xylene have been
shown to cause irreversible hearing loss in rats, Gagnaire and Langlais (2005) tested the relative
ototoxicity of 21 aromatic solvents, including //-butylbenzene. In their studies, groups of 7-8 young
3

-------
FINAL
9-13-2010
male Sprague-Dawley rats were administered 8.47 mmol/kg of chemical (in a volume of 2 mL/kg) by
gastric intubation for 5 days/week for a 2-week period2. Using the molecular weight of 134.22 g/mol
for //-butylbenzene, a molar concentration of 8.47 mmol/kg is equivalent to a dose of 1137 mg/kg-day.
After dosing, body weights were measured daily during the 2 weeks of treatment, and then for a
subsequent 10 days after the period of treatment. The behavior and general health of rats was observed
on a daily basis. At the end of the 10-day recovery period, six rats per treatment group were chosen
randomly, deeply anesthetized, and perfused with buffered paraformaldehyde and glutaraldehyde.
Subsequently, three left and three right cochleas were removed from the six chosen rats in each group
and processed. Organs of Corti and basilar membranes were examined by light microscopy and
scanning electron microscopy.
The only mortality was observed in 2/8 rats treated with isobutylbenzene (Gagnaire and
Langlais, 2005). The study authors noted ataxia and hypoactivity in the rats treated with
isobutylbenzene after each treatment. No treatment-related clinical signs were observed in any of the
other groups—including those treated with //-butylbenzene. Of the 21 solvents tested, the following
eight caused histological lesions (loss of hair cells) in the organ of Corti (listed from most to least toxic
based on cytocochleograms3): allylbenzene, ethylbenzene, styrene, //-propylbenzene, /^-xylene, toluene,
/ra//.v-P-methyl styrene, and a-methylstyrene. The remaining chemicals tested, including cumene
(isopropylbenzene), ^-butylbenzene, /^/'/-butylbenzene, 1,4-diethylbenzene, sec-butylbenzene,
/>ethyltoluene, 2-,3- and 4-methylstyrene, w-xylene, o-xylene, and benzene did not cause biologically
significant inner or outer hair cell loss and were considered to be inactive with regard to ototoxicity.
Following an examination of octanol/water partition coefficients for the chemicals tested, Gagnaire and
Langlais (2005) concluded that there was no correlation between ototoxicity and lipophilicity and that an
unidentified structural constraint was essential to induce ototoxicity. Given that only one dose was
tested, a freestanding NOAEL of 1137 mg/kg-day is identified for ^-butylbenzene in this study. For
comparative purposes, the LOAELs for ethylbenzene and //-propylbenzene in this study are 899 and
1018 mg/kg-day, respectively, based on molecular weights of 106.16 and 120.19 g/mol, respectively.
The RD50 (concentration necessary to depress the respiratory rate by 50% during acute exposure)
for sensory irritation by ^-butylbenzene was 710 ppm in a 30-minute exposure; the chemical did not
cause pulmonary irritation (defined as a decrease in respiratory rate during exposure via tracheal
cannula) at the RD50 (Nielsen and Alarie, 1982).
DERIVATION OF PROVISIONAL SUBCHRONIC AND CHRONIC
ORAL RfD VALUES FOR /i-BUTYLBENZENE
SUBCHRONIC p-RfD
There are no subchronic systemic toxicity studies of //-butylbenzene. The two-generation
reproduction study of rats by Izumi et al. (2005) is comprehensive and well conducted and addresses
2The dose was selected on the basis of previous range-finding studies conducted with toluene. The chosen dose was
associated with outer hair cell (OHC) loss in the middle turn of the organ of Corti—without causing mortality or body-weight
loss.
3Cytocochleograms are three-dimensional graphs based on counts of the inner hair cells (IHC) and three rows of OHC in the
organ of Corti.
4

-------
FINAL
9-13-2010
variables relevant to neurotoxicity and endocrine disruption as well as the usual spectrum of variables
typically assessed in a multigeneration reproduction study. It is the chosen principal study. Statistically
significant increases (p < 0.05) in organ weights were observed (liver, kidneys, and adrenals in parental
animals of two generations; and thymus weights in weanlings of F1 and F2 generations); only the
increased liver and kidney weights were supported by histopathological changes (hepatocellular
hypertrophy and hyaline droplets in proximal tubules, respectively) in F0 and F1 parent males. The
study authors reported observing no treatment-related effects on reproduction, reproduction hormones,
or growth and development of offspring over two generations (with exception of the increased thymus
weight).
Several effects were observed at 300 mg/kg-day in parental animals (increased liver, kidney and
adrenal weights, formation of hyaline droplet in proximal tubules, and hepatocellular hypertrophy), and
in F2 females (increased thymus weight). Although statistically significant changes in liver and kidney
weight were observed in parental animals at a dose of 100 mg/kg-day, these changes were considered
minimal and were not consistently seen across generations and sexes. In addition, there are concerns
about the significance of the thymus weight changes in F2 females such that this endpoint was not
chosen as the critical effect for the derivation of the p-subchronic RfD. Overall, the liver effects based
on the increased hepatocellular hypertrophy and liver weight are considered to be more sensitive than
the kidney effects, because these effects occurred in two generations (F0 and F1 parent males) and
increased liver weights were observed at lower doses (even though there were no histopathological
changes at 30- and 100-mg/kg-day treatment groups in either gender). Therefore, the critical effect is
hepatocellular hypertrophy with a LOAEL of 300 mg/kg-day in both F0 and F1 parent male rats.
To select a POD for subchronic p-RfD derivation, the increased incidences of hepatocellular
hypertrophy in F0 and F1 parent male rats (see Table 1) as the critical effect were modeled using EPA's
Benchmark Dose Software (v. 2.1). Appendix A provides details of the modeling effort and the
selection of the best fitting model. The best-fitting model, as assessed by AIC (model with lowest AIC)
for either data set was the gamma model. The BMDio and BMDLio derived by this model for the
F0 parent hepatocellular hypertrophy are 266 and 162 mg/kg-day, respectively. The BMDio and
BMDLio derived by this model for the F1 parent hepatocellular hypertrophy are 245 and 137 mg/kg-day,
respectively. The BMDLio of 137 mg/kg-day based on the F1 parent hepatocellular hypertrophy is
selected as the POD.
The subchronic p-RfD for «-butylbenzene is derived as follows:
Subchronic p-RfD = BMDLio UF
= 137 mg/kg-day ^ 1000
= 0.1 or 1 x 10"1 mg/kg-day
The composite UF of 1000 is composed of the following:
•	UFh: A factor of 10 is applied to account for intraspecies variability, including variability
in susceptibility in human populations and life-stages.
•	UFa: A factor of 10 is applied for animal-to-human extrapolation because data for
evaluating toxicokinetic or toxicodynamic differences are insufficient.
5

-------
FINAL
9-13-2010
•	UFd: A factor of 10 is applied for database inadequacies because neither general toxicity
or developmental studies are available. In addition, immunological toxicity is of
potential concern due to changes in thymus weight observed in F1 and F2 offspring.
•	UFl: A factor for extrapolating from a LOAEL to a NOAEL is not needed because BMD
modeling was used to determine the POD.
Confidence in the principal study is medium; although the study was well conducted and well
reported, it was not designed to address the full complement of variables normally addressed in a
subchronic toxicity study, and it does not include clinical chemistry, hematology, or urinalysis
components. Confidence in the database is low because it lacks true subchronic and developmental
toxicity studies. Thus, overall confidence in the subchronic p-RfD is low.
CHRONIC p-RfD
A chronic p-RfD is similarly derived by applying a UF of 3000 to the BMDLio of
137 mg/kg-day as follows:
Chronic p-RfD = BMDLio UF
= 137 mg/kg-day ^ 3000
= 0.05 or 5 x 10"2 mg/kg-day
The composite UF of 3000 is composed of the following:
•	UFh: A factor of 10 is applied to account for intraspecies variability, including variability
in susceptibility in human populations and life-stages.
•	UFa: A factor of 10 is applied for animal-to-human extrapolation because data for
evaluating toxicokinetic or toxicodynamic differences are insufficient.
•	UFd: A factor of 10 is applied for database inadequacies because neither general toxicity
or developmental studies are available. In addition, immunological toxicity is of
potential concern due to changes in thymus weight observed in F1 and F2 offspring.
•	UFl: A factor for extrapolating from a LOAEL to a NOAEL is not needed because BMD
modeling was used to determine the POD.
•	UFS: A factor of 3 is applied for using data from the two-generational reproductive study
(Izumi et al., 2005) based on the increased incidences of hepatocellular hypertrophy, in
both F0 and F1 parent males. The dose-response trends are similar in both F0 and F1
parent males, which suggest longer exposure {in utero and 18-week exposures) to
«-butylbenzene in F1 parent males may not lead to an increase in the incidences of
hypertrophy.
Confidence in the key study (Izumi et al., 2005) is medium, as discussed above for the
subchronic p-RfD. Confidence in the database for the chronic RfD is low due to the lack of subchronic,
chronic, and additional developmental toxicity studies. Thus, overall confidence in the chronic p-RfD is
low.
6

-------
FINAL
9-13-2010
FEASIBILITY OF DERIVING PROVISIONAL SUBCHRONIC AND CHRONIC
INHALATION RfC VALUES FOR /i-BUTYLBENZENE
Data on the inhalation toxicity of the //-butylbenzene are limited to an acute respiratory irritation
study that is not appropriate as the basis for the derivation of provisional RfCs.
PROVISIONAL CARCINOGENICITY ASSESSMENT
FOR /f-BUTYLBENZENE
WEIGHT-OF-EVIDENCE DESCRIPTOR
Under the 2005 Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005), there is
"Inadequate Information to Assess Carcinogenic Potential" of //-butylbenzene. There are no human
epidemiology studies, genotoxicity studies, or carcinogenicity assays.
QUANTITATIVE ESTIMATES OF CARCINOGENIC RISK
The lack of data on the carcinogenicity of ^-butylbenzene precludes the derivation of
quantitative estimates of risk for either oral or inhalation exposure.
REFERENCES
ACGIH (American Conference of Governmental Industrial Hygienists). 2007. 2007 Threshold Limit
Values for Chemical Substances and Physical Agents and Biological Exposure Indices. ACGIH,
Cincinnati, OH.
ATSDR (Agency for Toxic Substances and Disease Registry). 2008. Toxicological Profile Information
Sheet. U.S. Department of Health and Human Services, Public Health Service. Online,
http ://www. atsdr. cdc.gov/toxpro2. html.
Gagnaire, F. and C. Langlais. 2005. Relative ototoxicity of 21 aromatic solvents. Arch. Toxicol.
79(6):346-354.
Gerarde, H.W. 1959. Toxicological studies on hydrocarbons. III. The biochemorphology of the
phenylalkanes and phenylalkenes. AMA Arch. Ind. Health. 19:403-418.
Haley, P.J. 2003. Species differences in the structure and function of the immune system. Toxicology.
188:49-71.
7

-------
FINAL
9-13-2010
Henderson, R.F. 2001. Aromatic Hydrocarbons-Benzene and Other Akylbenzenes. In: Patty's
Toxicology. 5th Ed. E. Bingham, B. Cohrssen and C.H. Powel, Eds. John Wiley and Sons, New York.
4:231-301.
IARC (International Agency for Research on Cancer). 2008. Search IARC Monographs. Online.
http://monographs.iarc.fr/ENG/Monographs/allmonos90.php.
Imaoka, S. and Y. Funane. 1991. Induction of cytochrome P450 isoenzymes in rat liver by methyl n-
alkyl ketones and //-alkylbenzenes. Effects of hydrophobicity of inducers on inducibility of cytochrome
P450. Biochemical Pharmacology, 42 (Suppl.): S143-S150.
Izumi, H., Kimura, E., Ota, T., and Shimazu, S. 2005. A two-generation reproductive toxicity study of
n-butylbenzene in rats. The Journal of Toxicological Sciences, 30 (Special Issue):
21-38.
Nielsen, G.D. and Y. Alarie. 1982. Sensory irritation, pulmonary irritation, and respiratory stimulation
by airborne benzene and alkylbenzenes: prediction of safe industrial exposure levels and correlation
with their thermodynamic properties. Toxicol. Appl. Pharmacol. 65:459-477.
NIOSH (National Institute for Occupational Safety and Health). 2008. NIOSH Pocket Guide to
Chemical Hazards. Index by CASRN. Online, http://www2.cdc.gov/nioshtic-2/nioshtic2.htm.
NTP (National Toxicology Program). 2005. 11th Report on Carcinogens. U.S. Department of Health
and Human Services, Public Health Service, National Institutes of Health, Research Triangle Park, NC.
Online, http://ntp-server.niehs.nih.gov.
NTP (National Toxicology Program). 2008. Management Status Report. Online.
http://ntp.niehs.nih.gov/index.cfm?objectid=78CC7E4C-FlF6-975E-72940974DE301C3F.
OSHA (Occupational Safety and Health Administration). 2008. OSHA Standard 1910.1000 Table Z-l.
Part Z, Toxic and Hazardous Substances. Online, http://www.osha-
slc.gov/OshStd_data/l 910 1000 TABLEZ-1 .html.
Tanii, H., J. Huang and K. Hashimoto. 1995. Structure-Acute Toxicity Relationship of
Aromatic Hydrocarbons in Mice. Toxicol. Lett. 76:27-31.
U.S. EPA. 1987. Drinking Water Health Advisory for n-Butylbenzene. Environmental Criteria and
Assessment Office. Cincinnati, OH. Office of Health and Environmental Assessment, Cincinnati, OH.
U.S. EPA. 1991. Chemical Assessments and Related Activities (CARA). Office of Health and
Environmental Assessment, Washington, DC.
U.S. EPA. 1994. Chemical Assessments and Related Activities (CARA). Office of Health and
Environmental Assessment, Washington, DC. December.
8

-------
FINAL
9-13-2010
U.S. EPA. 1997. Health Effects Assessment Summary Tables. FY-1997 Update. Prepared by the
Office of Research and Development, National Center for Environmental Assessment, Cincinnati OH
for the Office of Emergency and Remedial Response, Washington, DC. July. EPA/540/R-97/036.
NTIS PB97-921199.
U.S. EPA. 2000. Benchmark Dose Technical Guidance Document. Risk Assessment Forum,
Washington, DC. External Review Draft. EPA/630/R-00/001.
U.S. EPA. 2005. 2005. Guidelines for carcinogen risk assessment. Risk Assessment Forum,
Washington, DC; EPA/630/P-03/001F. Federal Register 70(66):17765-17817
U.S. EPA. 2006. 2006 Edition of the Drinking Water Standards and Health Advisories. Office of
Water, Washington, DC. EPA 822-R-06-013. Washington, DC. Online.
http://www.epa.gov/waterscience/drinking/standards/dwstandards.pdf.
U.S. EPA. 2008. Integrated Risk Information System (IRIS). Office of Research and Development,
National Center for Environmental Assessment, Washington, DC. Online, http://www.epa.gov/iris/.
WHO (World Health Organization). 2008. Online catalogs for the Environmental Health Criteria
Series. Online, http://www.who.int/ipcs/publications/ehc/ehc_alphabetical/en/index.html.
Yamasaki, K., T. Michihito and M. Yasuda. 2005. Two-generation Reproductive Toxicity Studies in
Rats with Extra Parameters for Detecting Endocrine Disrupting Activity: Introductory Overview of
Results for Nine Chemicals. J.Toxicol. Sci. 30:1-4.
9

-------
FINAL
9-13-2010
APPENDIX A. DETAILS OF BENCHMARK DOSE MODELING
FOR SUBCHRONIC AND CHRONIC p-RfDs
MODEL FITTING PROCEDURE FOR QUANTAL NONCANCER DATA
The model-fitting procedure for dichotomous noncancer data is as follows. All available
dichotomous models in the EPA Benchmark Dose Software (BMDS, version 2.1) are fit to the incidence
data using the extra risk option. The multistage model is run for all polynomial degrees up to n-1 (where
n is the number of dose groups including control). Goodness-of-fit is assessed by the % test. When
several models provide adequate fit to the data (% p >_0.1), and the estimated BMDLs from these
models differ by >3-fold, then the model with the lowest BMDL is selected. Otherwise, models with
adequate fit are compared using the Akaike Information Criterion (AIC). The model with the lowest
AIC is considered to provide the best fit to the data. When several models have the same AIC, the
model resulting in the lowest BMDL is selected. In accordance with U.S. EPA (2000) guidance,
benchmark doses (BMDs) and lower bounds on the BMD (BMDLs) associated with an extra risk of
10% (BMDio and BMDLio) are calculated for all models.
Model-Fitting Results for Liver Hypertrophy in F0 and F1 Males (Izumi et al., 2005)
Applying the procedure outlined above to the F0 and F1 male data on the incidences of
hepatocellular hypertrophy (see Table 1), an adequate model fit was achieved with several models for
both data sets. Table A-l and A-2 show the results for the liver effect. In accordance with U.S. EPA
(2000) guidance, the model with the lowest AIC was considered to provide the best fit to the data. For
the F0 males, the resulting benchmark dose (BMDio) and associated 95% lower confidence limit
(BMDLio) are 266 and 162 mg/kg-day, respectively. For the F1 males, the resulting benchmark dose
(BMDio) and associated 95% lower confidence limit (BMDLio) are 245 and 137 mg/kg-day,
respectively. Figure A-l shows the model fit of the Gamma model to the F0 data; this model results in
the lowest AIC value and best fit to the data. Figure A-2 shows the model fit of the Gamma model to
the F1 data; this model results in the lowest AIC value and best fit to the data.
10

-------
FINAL
9-13-2010
Table A-l. Benchmark Dose Model Predictions for Liver Hypertrophy in
F0 Male Ratsa
Model
Degrees of
Freedom
x2
X2 Goodness
of Fit
/7-Value
AIC
BMD10
(mg/kg-day)
BMDL10
(mg/kg-day)
Quantal Linear
3
2.44
0.49
30.52
201.25
104.39
Multistage (degree = l)b
3
2.44
0.49
30.52
201.25
104.39
Multistage (degree = 2)b
3
0.68
0.88
27.84
214.77
148.71
Multistage (degree = 3)b
3
0.68
0.88
27.84
214.77
148.71
Weibull (power > 1)
2
0
1
28.56
285.42
164.82
Gamma (power > 1)
3
0
1
26.56
265.73
161.56
Probit
2
0
1
28.56
280.75
197.65
Log-probit (slope >1)
2
0
1
28.56
270.06
153.76
Log-logistic (slope >1)
2
0
1
28.56
284.65
161.76
Logistic
2
0
1
28.56
290.25
211.73
aIzumi et al., 2005.
bDegree of polynomial initially set to (n - 1) where n = number of dose groups including control. Betas restricted
to >0.
Table A-2. Benchmark Dose Model Predictions for Liver Hypertrophy in
F1 Male Ratsa
Model
Degrees of
Freedom
x2
X2 Goodness
of Fit
/7-Value
AIC
BMD10
(mg/kg-day)
BMDL10
(mg/kg-day)
Quantal Linear
3
3.37
0.3376
31.05
130.54
71.21
Multistage (degree = l)b
3
3.37
0.338
31.05
130.54
71.21
Multistage (degree = 2)b
3
0.97
0.81
27.47
170.34
115.86
Multistage (degree = 3)b
3
0.97
0.81
27.47
170.34
115.86
Weibull (power > 1)
2
0
1
27.70
277.34
140.05
Gamma (power > 1)
3
0
1
25.70
244.52
136.74
Probit
2
0
1
27.70
269.29
169.90
Log-probit (slope >1)
2
0
1
27.70
253.87
130.53
Log-logistic (slope >1)
2
0
1
27.70
275.43
136.95
Logistic
2
0
1
27.70
284.28
184.69
aIzumi et al., 2005.
bDegree of polynomial initially set to (n - 1) where n = number of dose groups including control. Betas restricted
to >0.
11

-------
FINAL
9-13-2010
Gamma Multi-Hit Model with 0.95 Confidence Level
0.4
0.3
~o
Q)
O
<
.1	0.2
o
ra
0.1
0
0	50	100	150	200	250	300
Dose
17:20 10/05 2009
Figure A-l. Fit of Gamma Model to Data on Hepatocellular Hypertrophy in F0 Male Rats
BMDs and BMDLs indicated are associated with an extra risk of 10% and are in units of mg/kg-day
Gamma Model. (Version: 2.13; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS21Beta\Temp\ltmpl45B.(d)
Gnuplot Plotting File: C:\USEPA\BMDS21Beta\Temp\ltmpl45B.plt
Mon Oct 05 17:20:24 2009
BMDS Model Run
The form of the probability function is:
P[response]= background+(1-background)*CumGamma[slope*dose, power],
where CumGamma(.) is the cummulative Gamma distribution function
Dependent variable = Incidence
Independent variable = Dose
Power parameter is restricted as power >=1
Total number of observations = 4
Total number of records with missing values = 0
Maximum number of iterations = 25 0
Relative Function Convergence has been set to: le-008
12

-------
FINAL
9-13-2010
Parameter Convergence has been set to: le-008
Default Initial	(and Specified) Parameter Values
Background =	0.02
Slope =	0.00414503
Power =	2.6395 6
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Background -Power
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Slope
Slope	1
Parameter Estimates
Variable
Background
Slope
Power
Estimate
0
0.0482501
18
Std. Err.
NA
0. 0035348
NA
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
0.041322
0. 0551782
NA - Indicates that this parameter has hit a bound
implied by some ineguality constraint and thus
has no standard error.
Analysis of Deviance Table
Model
Full model
Fitted model
Reduced model
Log(likelihood)
-12 .2818
-12.2818
-19.642
# Param's Deviance	Test d.f.
4
1 0.000161175	3
1 14.7205	3
P-value
0. 002072
AIC:
26.5637
Dose
Goodness of Fit
Est. Prob.
Expected
Observed
Size
Scaled
Residual
0. 0000
30.0000
100.0000
300.0000
Chi^2 =0.00
0.0000
0.0000
0.0000
0.2083
d.f.
0.000	0.000	24
0.000	0.000	24
0.000	0.000	24
5.000	5.000	24
P-value = 1.0000
0.000
-0.000
-0.009
0.000
Benchmark Dose Computation
Specified effect =	0.1
Risk Type	=	Extra risk
13

-------
Confidence level
BMD
BMDL
FINAL
9-13-2010
0.95
265.733
161.563
14

-------
FINAL
9-13-2010
Gamma Multi-Hit Model with 0.95 Confidence Level
Dose
15:52 09/21 2009
Figure A-2. Fit of Gamma Model to Data on Hepatocellular Hypertrophy in F1 Male Rats
BMDs and BMDLs indicated are associated with an extra risk of 10% and are in units of mg/kg-day
Gamma Model. (Version: 2.13; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS21Beta\Temp\ltmpl3B0.(d)
Gnuplot Plotting File: C:\USEPA\BMDS21Beta\Temp\ltmpl3B0.plt
Mon Sep 21 15:46:54 2009
BMDS Model Run
The form of the probability function is:
P[response]= background+(1-background)*CumGamma[slope*dose, power],
where CumGamma(.) is the cummulative Gamma distribution function
Dependent variable = Incidence
Independent variable = Dose
Power parameter is restricted as power >=1
15

-------
FINAL
9-13-2010
Total number of observations = 4
Total number of records with missing values = 0
Maximum number of iterations = 25 0
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial	(and Specified) Parameter Values
Background =	0.025
Slope =	0.00622162
Power =	2.9511
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Background -Power
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Slope
Slope	1
Parameter Estimates
Variable
Background
Slope
Power
Estimate
0
0.0524367
18
Std. Err.
NA
0.00388009
NA
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
0. 0448319
0.0600415
NA - Indicates that this parameter has hit a bound
implied by some ineguality constraint and thus
has no standard error.
Analysis of Deviance Table
Model
Full model
Fitted model
Reduced model
Log(likelihood)
-11.8494
-11.8497
-21.1528
# Param's
4
1
1
Deviance Test d.f.
0.000427008
18.6067
P-value
0.0003297
AIC:
25.6993
Dose
Est. Prob.
Goodness of Fit
Expected
Observed
Size
Scaled
Residual
0. 0000
30.0000
100.0000
300.0000
0.0000
0.0000
0.0000
0.3158
0. 000
0. 000
0. 000
6. 000
0. 000
0. 000
0. 000
6. 000
19
19
21
19
0.000
-0.000
-0.015
0.000
Chi^2 =0.00
d.f.
P-value = 1.0000
Benchmark Dose Computation
16

-------
Specified effect
Risk Type
Confidence level
BMD
BMDL
FINAL
9-13-2010
o.i
Extra risk
0.95
244.517
136.737
17

-------