Scientific Review of 1-Bromopropane Occupational Exposure Limit 

Derivations – 

Preliminary Thoughts and 

Areas for Further Analysis

Scientific Review of 1-Bromopropane Occupational Exposure Limit
Derivations - Preliminary Thoughts and Areas for Further Analysis

Abstract

Current OELs for 1-bromopropane (1-BP) are diverse in both the selection
of critical effects and judgments of remaining uncertainties.  The
resulting values differ by ~16-fold.  We critically evaluated the
underlying basis of existing OELs through the use of concepts such as
critical effect, benchmark dose and uncertainty factor.  We conclude
that the critical effect is decreased live litter size with a BMDL of
190 ppm.  Using an uncertainty factor of 10-fold, 3 for extrapolation
from an animal study and 3 for human variability results in an OEL of 20
ppm.

Introduction

The development of Occupational Exposure Limits (OELs) for various
chemicals found in our workspace is an important endeavor for risk
assessment scientists and managers.  1-Bromopropane (1-BP) is used as a
solvent for fats, waxes, or resins, as an intermediate in the synthesis
of numerous products, including pharmaceuticals, insecticides, flavors
and fragrances, and as a solvent in spray adhesives and as a degreaser. 
Current OEL derivations for 1-BP include a number of organizations or
investigators, presented in Table 1.

As shown in Table 1, current OELs for 1-BP are diverse in both the
selection of critical effects and judgments of remaining uncertainties. 
The resulting OELs differ by ~16-fold.  Some of these differences
reflect the year of evaluation and lack of recently published studies. 
The purpose of this paper is to critically evaluate the underlying basis
of existing OELs and bring some common understanding through the use of
concepts such as critical effect, benchmark dose and uncertainty factor.
 We conclude that the most appropriate OEL for 1-BP takes from elements
of each of these current OELs.

Table 1.  OELs Derived for Various Groups and Their Basis.

Group	OEL (ppm)	Critical Effect	Uncertainty	Reference

ACGIH TLV (2004) 

	10	LOAEL of 100 ppm for decreased fetal weight	An apparent factor of 10
was used, no specific factor or rationale was provided	Huntingdon, 2001

Stelljes and Wood (2004)	156	BMDL of 156 ppm for decreased sperm
motility in F1 generation	1-fold, no uncertainty remains	WIL Research,
2001

Rozman and Doull (2002)	60 - 90	NOAEL of 170 ppm for mild CNS effects
(headache) in workers	2-3-fold for within human variability	NIOSH, 2000

U.S. EPA (2002)	20	BMDL  (adjusted) of 177 ppm for decreased sperm
motility in F1 generation	10-fold for within human and animal to human
variability	WIL Research, 2001

ICF (1998)

	100	NOAEL (adjusted) of 300 ppm for mild liver histopathology and NOAEL
of 280 ppm for decreased sperm motility	3-fold for animal to human
variability	ClinTrials BioResearch, 1997



Methods

The methods used in this review are those as published by Haber et al.
(2001).  In brief, we use the concept of critical effect, benchmark dose
(BMD) and uncertainty factor as described by these authors because we
feel that such concepts can be useful in the determination of OELs in
general, and specifically can be used to harmonize seemingly disparate
judgments for 1-BP.

The critical effect is defined as the first adverse effect, or its known
precursor, that occurs as dose rate or concentration increases.  One or
more effects may be critical for any particular chemical.  This concept
is used world wide for both environmental and occupational risk
assessments. 

Uncertainty factors are considered a necessary reduction in the exposure
level, based on scientific judgments of available toxicity,
toxicokinetics or toxicodynamics and inherent uncertainty.  Although
default values of 10-fold are commonly used for different areas of
uncertainty, especially in environmental risk assessment, such defaults
are seldom used in occupational risk assessment.  However, the
environmental area of assessment is now emphasizing the use of specific
data and better judgment in the development of uncertainty factors,
rather than the usual default values of 10-fold (Dourson et al., 1996;
IPCS, 2001).  The occupational area of assessment is now emphasizing a
more structured approach to uncertainty judgments (Naumann and Weidman,
1995; Haber and Maier, 2002).  Thus, we feel that the two lines of
judgment are not as far apart as some scientists might think.

The benchmark dose (BMD) approach was also used in conjunction with the
more standard NOAEL/LOAEL technique to analyze the data for 1-BP.  The
use of both approaches necessitates expert judgment and adds value to
the overall assessment.  U.S. EPA’s BMD software, version 1.3.2 (U.S.
EPA, 2001) was used to reproduce each critical benchmark dose (BMD) and
lower bound benchmark dose (BMDL) for 1-BP calculated by Stelljes and
Wood (2004) and to expand the analysis to additional endpoints of
interest for the 2-generation study (WIL Research, 2001).  Benchmark
responses (BMRs) of 1.0 control standard deviation were used by TERA for
all continuous data and BMRs of 10% were used for all dichotomous data. 
These choices reflect standard operating procedure.  All the available
models in the BMDS software were run for each data set, and BMDs and
BMDLs from the best fitting model were selected.

Results and Discussion

Issues related to Critical Effect with Existing OEL Estimates

Neurotoxicity.  Neurotoxicity is a common effect from exposure to 1-BP. 
However, its selection as a critical effect is made difficult because of
inconsistencies in the overall database.  For example, an argument is
presented by Stelljes and Wood (2004) that CNS vacuolization found in
the Ichihara et al. (2000a) study should not serve as the critical
effect.  This argument needs to be sharpened, particularly with regard
to the absence of the finding in longer-term studies and the possibility
that the effect resulted from methods used for tissue preparation. 
However, this argument is consistent with the fact that none of the
longer-term studies, including preliminary data from the 13-week NTP
study (NTP, 2003), identified CNS histopathology changes.  

The finding of reduced hind-limb grip strength from the Ichihara et al.
(2000a) study would generally be an appropriate endpoint for risk
assessment with a BMDL calculated by Stelljes and Wood (2004) of 214 ppm
(Table 2).  However, documented concerns with the conduct of the study
reporting this finding (O’Malley, 1999), the inability of GLP 90-day
study (ClinTrials BioResearch, 1997) to duplicate this finding, the
absence of CNS histopathology in the NTP 13-week study (NTP, 2003), and
inconclusive evidence of psychomotor performance effects in an
investigation of workers (NIOSH, 2002) weaken the argument that the BMDL
of 214 ppm for reduced hind-limb strength is an appropriate choice for
the critical effect level.  

Neurological effects remain of interest based on recent neurotoxicity
studies that suggest spontaneous locomotor activity was raised in rats
at exposure of 50 ppm and higher; other clinical signs were also noted
as statistically significant at exposures of 200 or 1000 ppm (Wang et
al., 2003).  Biochemical changes in the brain of rats occurred at
exposures of 200 ppm and higher (Honma et al., 2003).  Subjective
symptoms were also reported in human case studies at average exposures
of about 60 to 70 ppm, but the effects were not definitively
attributable to 1-BP (NIOSH, 2002), or an increased incidence of
headaches at average concentrations between 190 and 200 ppm (NIOSH,
2000).  This latter effect was the basis of the OEL proposed by Rozman
and Doull (2002).

For the neurotoxicity findings the bottom line appears to be that some
animal and human studies suggest effects in the range of 100 to 200 ppm
or higher, but results across the overall database are not consistent. 
Furthermore, definitive effect levels in these studies generally fall in
the same range as for reproductive toxicity endpoints (which are of
greater severity as shown below).  The human data are limited by
co-exposure to other solvents, small populations examined, and
limitations in the exposure estimates.  Due to these uncertainties, the
human data appear inadequate to serve as the primary basis for the
critical effect as suggested by Rozman and Doull (2002), although they
are quite useful in serving as a comparison to any derived OEL.

Liver Toxicity.  Most risk assessors would consider the increased
incidence of mild liver cytoplasmic vacuolization, such as that seen by
ClinTrials BioResearch (1997), as a minimal adverse effect, even though
other measures of liver damage (e.g., serum levels of liver enzymes)
were not affected in this study.  In fact, the lack of additional liver
effects further supports the mild cytoplasmic vacuolization as an effect
of minimal severity.  The BMDL of 226 ppm determined by Stelljes and
Wood (2004) for this endpoint corresponds well with the effect level for
other endpoints, although the severity is minimal (Table 2). 
Furthermore, no severe treatment related liver findings (histopathology
or clinical chemistry changes) were reported in preliminary data from
the 13-week NTP study (NTP, 2003).  These data suggest that liver
toxicity is not the critical target for 1-BP toxicity, and the older ICF
(1998) OEL should be discounted.

Reproductive and Fetal Effects.  The ACGIH (2004) use of a LOAEL of 100
ppm for decreased fetal weight in the Huntingdon Study (2001) as the
critical basis for its OEL derivation is difficult to justify.  As a
first consideration, BMDL estimates for this endpoint are greater than
for other effects.  For example, both the NTP expert panel (NTP, 2002)
and TERA (shown later in this text and in Table 2) identified a BMDL of
approximately 300 ppm.  Furthermore, questions about the conduct of the
Huntingdon (2001) study, including the change in procedure with control
animals that lead to higher body weights, lack of related findings of
developmental delays in pups in multi-generation studies at similar
concentrations (see WIL Research, 2001), minimal severity of the effect
(a maximum of 7% change from control), and potential relatedness to
maternal effects (although BMDL for pup fetal weight is lower than for
maternal weight) decreases the selection of this endpoint as the most
relevant for deriving the OEL.  

Stelljes and Wood (2004) argue that the effect level for sperm
parameters in the WIL Research (2001) study should be based on the F0
generation results and not those for F1 or F2 animals, because the goal
of an OEL is to develop a safe exposure level for workers and the
exposure patterns for the parental animals more closely resemble
occupational exposure scenarios.  A counter to this argument is that in
utero exposure may cause effects manifested as these exposed animals
become adult males.  However, it is not clear what mechanisms would
generate changes in sperm parameters based on the normal turnover in
sperm through the cycle of spermatogenesis, in the absence of findings
on male reproductive organ histopathology.  The BMDL calculated by
Stelljes and Wood (2004) for sperm motility in F0 animals was 263 ppm. 
TERA identified a BMDL of 270 ppm (see later in this text).  Based on
data from Ichihara et al. (2000b), Stelljes and Wood (2004) calculated a
BMDL of 232 ppm for sperm count in F0 adult males.  Taken together, the
effects of 1-BP on male sperm parameters suggests that the male
reproductive effect in parental animals occur in the same general range,
but are not more sensitive than other relevant effects.

Benchmark dose modeling for several measures of male and female
reproductive parameters from the two-generation study correspond well
with each other and provide a consistent story indicating that 1-BP can
affect reproductive parameters in males (decreased sperm motility and
prostate weight), females (increased estrous cycle length, no estrous
cycle incidence, and maternal body weight at gestation day 20), and
functional reproductive performance (litter viability index, pup weight
gain at post natal days 21 to 29, and live litter size).  The BMDL
values for these latter effects are in the same range, but slightly
lower, than for the liver effects, and represent a more serious outcome
(Table 2).  The BMDL value of 188 ppm from Stelljes and Wood (2004) or
of 190 ppm from TERA for decreased live F1 litter size is the most
appropriate basis for deriving the OEL, since this is the lowest measure
related to exposure to F0 animals that is clearly adverse. 

Issues related to Critical Effect with New Studies

A 90-day inhalation study (NTP, 2003) was conducted in rats and mice. 
Male and female B6C3F1 mice were exposed to 0, 62.5, 125, 250, and 500
ppm 1-BP for 90 days. Male and female Fischer 344 rats were similarly
exposed to 0, 62.5, 125, 250, 500, and 1000 ppm for 90 days. 

Body weight gain and terminal body weights of female and male mice and
female rats were comparable to controls.  In rats given 1000 ppm, there
was reduction in body weight gain and terminal body weight, which became
significant beginning at week 9.  No mortality was observed in rats.  In
mice, no mortality was observed in animals given 250 ppm or below but
3/10 female and 2/10 male animals died from natural causes and 2 females
and 2 males found moribund were killed.  The only clinical signs of
toxicity observed included abnormal breathing and lethargy in 2 male and
2 female mice.  No significant signs of clinical toxicity were observed
in the rats.  It further appeared that 1-BP did not cause any adverse
effects on clinical chemistry parameters.  There were some slight,
dose-dependent but insignificant changes in some of the parameters but
these changes are not likely to be of toxicological significance. 
Hematological parameters were also not adversely affected in mice or
rats. 

Microscopic evaluation revealed no significant abnormality at 250 ppm or
below.  At 500 ppm, mice that died or were killed in extremis or
survived to the end of the study period had mild, chronic inflammation,
marked necrosis, and mild cytoplasmic vacuolization in the centrilobular
hepatocytes, moderate to marked necrosis of the adrenal cortex, mild to
moderate necrosis and moderate cytoplasmic vacuolization of the
bronchioles and trachea, and minimal to moderate necrosis and
cytoplasmic vacuolization of the trachea. Similar observations were
noted in female and male rats at 500 ppm and 1000 ppm. Minimal
cytoplasmic vacuolization of the centrilobular hepatocytes were observed
at 250 ppm, the severity of which increased at higher doses.  Based on
these results, it appears that the body weight changes observed in male
rats at 1000 ppm were accompanied by microscopic abnormalities,
indicating a possible LOAEL of 1000 ppm and a NOAEL of 500 ppm.  In
mice, the frank toxicity at 500 ppm was also accompanied by microscopic
abnormalities, indicating a possible LOAEL of 500 ppm and a NOAEL of 250
ppm. 

Wang et al. (2003) reported decreased creatinine kinase activities in
central nervous system tissues following 12-week exposures to
concentrations beginning at 200 ppm in rats.  This is the same group
that published the Ichihara et al. (2000a) study, and reports a fairly
obscure endpoint for risk assessment purposes.   

Honma et al. (2003) reported decreased body weight at 1000 ppm
(consistent with the NTP (2003) study), effects on locomotor activity at
50 and 200 ppm (although this was measured as latency in recovery of
activity), changes in ambulation and rearing at 200 ppm (but not 1000
ppm), and changes in performance in a traction test (at 200 and 1000
ppm).  TERA has not had sufficient time to closely analyze this study,
and would welcome ACGIH thoughts.  

In summary, NOAEL/LOAEL and BMD and BMDL boundaries for experimental
animal male and female nervous system, liver toxicity and reproductive
and fetal effects are in the same range, suggesting that regardless of
endpoint selected the critical effect levels will not vary greatly. 
These boundaries are generally similar to that seen in humans.  This
increases confidence the overall OEL value derived will provide adequate
coverage for the range of potential endpoints.

Benchmark Dose Modeling for 1-BP

The results of Benchmark Dose (BMD) modeling are summarized in Table 2. 
TERA’s calculations were generally consistent with those reported by
Stelljes and Wood (2004).  For example, the BMDL of 263 ppm for F0 sperm
motility from Stelljes and Wood (2004) was similar to that of TERA of
270 ppm.  The BMD and BMDL computed by Stelljes and Wood (2004) for
centrilobular vacuolization of 345 and 226 ppm are consistent with the
results of using a multistage model of order 4.  However, the simpler
multistage model of order 2, which is commonly used in BMD modeling, was
used to recalculate a BMDL because the 2nd order model has a slightly
lower Akaike's Information Criterion (AIC) (38 versus 39) indicating a
superior data fit (p = 0.9) with fewer parameters.

Additional reproductive and developmental effects not considered by
Stelljes and Wood (2004) were also evaluated by TERA.  These endpoints
include F0 prostate weights, F0 and F1 estrous cycle lengths, number of
F0 and F1 rats not having estrous cycles, maternal body weight at
gestation day 20, F1 litter viability index, F1 pup weight gain data
from WIL Research (2001) as well as fetal weight data from Huntingdon
Life Sciences (2001), and F1 and F2 live litter size.  BMDs and BMDLs
are shown for all of these endpoints except for the litter viability
index, which exhibited no clear dose-response.  The number of rats not
having estrous cycles was analyzed because the cycle length data
analysis necessarily omitted these animals that were experiencing a more
severe cycle delay that could not be quantified in terms of days.

 

A BMDL for fetal weight reduction was computed using the data collected
by Huntingdon Life Sciences (2001).  Following the NTP-CERHR expert
panel on reproductive and developmental toxicity of 1-BP (NTP, 2002),
one litter in the 100 ppm dose group was excluded because the average
fetal weight was more than 3 standard deviations from the dose group
mean.  A BMDL of 310 ppm was estimated using a BMR of 1.0 control
standard deviation from the control mean (Huntingdon Life Sciences,
2001).  This estimate is similar to the BMDL of 305 ppm estimated by the
NTP-CERHR expert panel (NTP, 2002), and they are higher than the BMDLs
for reproductive endpoints; therefore, fetal weight reduction is
unlikely to be the most sensitive effect.  This conclusion is consistent
with the findings of the NTP expert panel.

To further evaluate whether developmental effects were the most
sensitive basis for deriving an OEL, a BMDL for maternal weight change
was also computed by TERA using the data collected by Huntingdon Life
Sciences (2001).  Note that the maternal body weight was calculated as
maternal weight on GD20 subtracted by the litter weight at birth.  A
BMDL of 690 ppm was estimated using a BMR of 1.0 control standard
deviation from the control mean.  This estimate is higher than the BMDL
of 310 ppm estimated by TERA for fetal weight reduction, indicating that
the change of fetal weight might be due to direct fetal toxicity from
exposure to the compound rather than only a secondary effect from
maternal toxicity.  On the other hand, no consistent dose-related effect
on pup weights was observed in the two-generation study, decreasing
concern related to the decreased fetal weight finding.  Furthermore,
since the BMDL for decreased fetal weight was greater than for other
reproductive parameters from the two-generation study, this effect
should be adequately addressed by an OEL that protects against
reproductive effects.  

Finally, F1 and F2 live litter size was assessed.  BMD and BMDL values
of 280 and 188 for the F1 generation, were determined by Stelljes and
Wood (2004).  These values were confirmed by TERA where values of 280
and 190 are shown (Table 2).  Stelljes and Wood (2004) and TERA also
found similar BMD and BMDL values for the F2 generation, although these
values were lower than that for the F1.  This effect, decrease in live
litter size, is of sufficient severity to warrant its choice as the
critical effect.  Although other effects might occur at the same, or
slightly lower exposures, they are not as toxicologically significant. 
The choice of the BMD and BMDL values of 280 and 190 for the F1
generation, rather than lower values from the F2 generation reflects the
desire to replicate the likely exposure in a worker population. 
Specifically, it is not anticipated that any human will have the
exposure pattern of an F2 animal.  In contrast, the occupational
exposure pattern of an F1 animal might occur in humans.

Table 2

BMD and BMDL Estimates*









	Endpoint	Stelljes and Wood	TERA



 	BMD (ppm)	BMDL (ppm)	BMR	BMD (ppm)	BMDL (ppm)	Model	Variance

	 

 





Hindlimb strength	286	214	1 sd	290	210	Linear	Homogeneous

Minimal centrilobular vacuolization males	345	226	10%	290	200
Multistage-2

	Fetal body weight	 

1 sd	510	310	Poly-2	Non-homogeneous

F0 sperm motility	343	263	1 sd	380	270	Linear	Homogeneous

F1 sperm motility	261	156	1 sd	260	150	Power	Non-homogeneous

F0 prostate weight	 

1 sd	740	190	Power	Homogeneous

F0 Estrous Cycle Length	 

1 sd	290	210	Power	Non-homogeneous

F1 Estrous Cycle Length	 

1 sd	810	400	Linear	Non-homogeneous

F0 No Estrous Cycle Incidence	 

10%	670	480	Multistage-2

	F1 No Estrous Cycle Incidence	 

10%	360	180	Quantal Linear

	Maternal GD20 body weight	 

1 sd	1000	690	Linear	Homogeneous

F1 litter viability index	 

 No dose-response



F1 pup weight gain PND 21 to 28	 

1 sd	240	180	Linear	Homogeneous

F1 decreased live litter size	280	188	1 sd	280	190	Linear
Non-homogeneous

F2 decreased live litter size	238	169	1 sd	240	170	Linear
Non-homogeneous



*See text for additional details.

Areas of Uncertainty

Most organizations that establish OEL’s do not have documented
approaches for addressing areas of uncertainty, rather a professional
judgment approach is used (Haber and Maier, 2002).  In order to evaluate
potential OELs for 1-BP, we structure a discussion around the U.S.
EPA’s approach that describes five areas of uncertainty.  However, in
keeping with the existing OEL approach, we were not constrained to using
EPA’s defaults.  

Interspecies Variability (UFA).  This area accounts for the differences
that occur between experimental animals and humans and is composed of
subfactors for toxicokinetics (how the body distributes and metabolizes
the chemical) and toxicodynamics (how the body responds to the
chemical).  The use of these two considerations is standard practice in
the context of environmental risk assessment (Dourson et al., 1996), and
is gaining acceptance for assessing occupational risk (Naumann and
Weidman, 1995).  

Ideally, a quantitative comparison of the toxicant concentrations (e.g.
AUC or Cmax) in the target organ between animal species and humans would
allow interspecies variability in toxicokinetics to be calculated. 
However, for 1-BP the information available is not adequate to allow
such estimation.  An alternative is to calculate the human equivalent
concentrations (HEC) from the animal data based on the chemical’s
properties and physiological differences between the tested animal
species and humans.  This dosimetric adjustment generally provides a
better estimate of the target organ doses following inhalation exposure
than simply dividing the exposure assessment exposure by a default
uncertainty factor of 10-fold.  If the HEC is used, a toxicokinetic
subfactor for interspecies variability is generally not needed because
the expected toxicokinetic difference has been considered to some extent
in the HEC calculation.  If no information is available on the
quantitative differences in the organ response to the toxicant of
interest between animals and humans, then a default value of 3 for this
toxicodynamic difference is used in environmental assessments.  If data
are available to adequately describe this variability, then actual data
may be used to replace this default value as well (IPCS, 2001).  

For 1-BP, dosimetric adjustment to the HEC per EPA’s methods (see for
example ICF, 1998) support using a factor of 1 to account for species
differences in toxicokinetics.  Toxicodynamic differences, however, also
need to be addressed.  There appears to be general consistency in effect
levels among species for various toxic endpoints.  For example, mild CNS
effects in humans, as summarized by Rozman and Doull (2002), were
observed in a range generally similar to the BMDL for hindlimb grip
strength in rats (see Table 2) and several of the clinical findings of
Wang et al. (2003).  Nevertheless, because there is residual concern
about relative sensitivity to reproductive effects, and humans might be
expected to be more sensitive to reproductive parameters (based on less
excess reproductive capacity) a factor to account for toxicodymanic
differences appears appropriate.  For example, the in vivo dose-response
information in humans is scant, and therefore comparative sensitivities
of humans and animals are hard to define from the available data. 
Furthermore, in vitro bioassays are available for both human and animal
cell cultures, including human hepatocytes, mouse lymphoma and bone
marrow cells, but no data were obtained from experiments on reproductive
system tissues.  Moreover, since the critical effect is decreased live
litter size, identifying a suite of relevant in vitro studies that could
be used to compare animal and human responsive sensitivities would be
difficult to obtain without a better understanding of the underlying
mechanism of this effect.  Since the available data do not provide
sufficient information for a quantitative estimation of toxicodynamic
variation, a default subfactor of 3 is appropriate for this area of
uncertainty.  Additional studies investigating relative sensitivities to
reproductive effects of 1-BP would be helpful to address this area of
uncertainty.

Intraspecies Variability (UFH).  This factor accounts for the natural
differences that occur among human subpopulations and for the fact that
some individuals are more sensitive than the average population.  This
factor is also composed of two subfactors – one to account for
toxicokinetic differences and one to account for toxicodynamic
differences.  If no information is available on human variability, then
a default value of 10 is generally used in the context of environmental
exposures to the general population.  If adequate information is
available on either toxicokinetic of toxicodynamics variability, then
this information is used to develop estimates of variability from the
data (IPCS, 2001; Meek et al., 2001).  Unfortunately for 1-BP, no
quantitative information regarding human variability in terms of
toxicokinetics and toxicodynamics was identified, and therefore,
data-derived estimates of human variability cannot be calculated. 

However, for worker populations the degree of variability in
toxicokinetic or toxicodynamic variability is expected to be lower than
for the general population.  Since some degree of variability in
response would be expected even among the worker population, a reduced
factor of about 3-fold is generally judged to be reasonable.  This is
similar to what Rozman and Doull (2002) suggest. 

Extrapolation from an Effect Level (UFL).  A BMDL was used with the
critical effect.  Generally no additional factor is considered needed in
these situations.

Extrapolation from Less than Lifetime to Lifetime Exposure (UFS).  This
factor is not generally used by groups that establish OELs (Haber and
Maier, 2002).  The database for 1-BP lacks a completed chronic study,
and therefore the likelihood that effects would progress with longer
duration exposures needs further evaluation.  However, the critical
effect appears to be on a reproduction parameter and the critical study
evaluated the period of interest.  Moreover, workers have been exposed
to 1-BP for more than short term exposures and their results are
considered in all of these OEL estimations.  Thus, it does not appear to
us that a factor is needed for this area.

Adequacy of the Database (UFD).  This factor is not overtly used by
groups that establish OELs (Haber and Maier, 2002).  However, OEL
decisions routinely consider whether the overall body of literature
determines that the most sensitive effects have been evaluated.  For
1-BP in particular, reproductive toxicity and possibly neurotoxicity and
liver toxicity appear to be the most sensitive effects.  A decrease in
live litter size appears to be the critical effect.  We do not see the
need for a factor for this area of uncertainty.

Determination of OEL

We conclude that the critical effect for the purpose of developing an
OEL is decreased live litter size in the F1 generation, with a BMDL of
190 ppm as shown in Table 2.  Dividing this BMDL with an uncertainty
factor of 10-fold, which is composed of 3-fold for extrapolation from an
experimental animal study to humans for expected toxicodynamic
differences and 3-fold for expected human variability in toxicokinetics
and toxicodynamics within the worker population, results in an OEL of 20
ppm.  This OEL could be potentially lower if results in workers show
definitive reproductive or other toxicity at levels lower than about 100
ppm.  This OEL could be potentially higher if the expected reproductive
response in experimental animals is shown to be similar to humans and at
similar levels.

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NIOSH (National Institute of Occupational Safety and Health), 2002.
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NTP (National Toxicology Program), 2002. NTP-CERHR Expert Panel report
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Wang, H., Ichihara, G., Ito, H., Kato, K., Kitoh, J., Yamada, T., Yu,
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 TERA identified a BMDL of 200 ppm for this endpoint.

 Note, this study is not yet published nor peer-reviewed by NTP.  The
raw animal data are posted on the NTP website, and TERA developed these
conclusions from the available data.

NTP’s two-year bioassay is currently in progress.

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