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Chapter
VIII.
Quantification
of
Toxicological
Effects
The
quantification
of
toxicological
effects
of
a
chemical
consists
of
separate
assessments
of
noncarcinogenic
and
carcinogenic
health
effects.
Unless
otherwise
specified,
chemicals
which
do
not
produce
carcinogenic
effects
are
believed
to
have
a
threshold
dose
below
which
no
adverse,
noncarcinogenic
health
effects
occur,
while
carcinogens
are
assumed
to
act
without
a
threshold.

A.
Introduction
to
Methods
A.
1
Quantification
of
Noncarcinogenic
Effects
In
quantification
of
noncarcinogenic
effects,
a
Reference
Dose
(
RfD)
(
formerly
called
the
Acceptable
Daily
Intake
(
ADI))
is
calculated.
The
RfD
is
"
an
estimate
(
with
uncertainty
spanning
approximately
an
order­
of­
magnitude)
of
a
daily
exposure
to
the
human
population
(
including
sensitive
subgroups)
that
is
likely
to
be
without
appreciable
risk
of
deleterious
effects
over
a
lifetime"
(
U.
S.
EPA,
1993).
The
RfD
is
derived
from
a
no
observed
adverse
effect
level
(
NOAEL),
lowest
observed
adverse
effect
level
(
LOAEL),
or
a
NOAEL
surrogate
such
as
a
benchmark
dose
identified
from
a
subchronic
or
chronic
study,
and
divided
by
a
composite
uncertainty
factor(
s).
The
RfD
is
calculated
as
follows:

RfD
=
NOAEL
(
LOAEL)
UF
×
MF
where:

NOAEL
=
No­
observed­
adverse­
effect
level
from
a
high­
quality
toxicological
study
of
an
appropriate
duration
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LOAEL
=
Lowest­
observed­
adverse­
effect
level
from
a
high­
quality
toxicological
study
of
an
appropriate
duration.
In
situations
where
there
is
no
NOAEL
for
a
contaminant
but
there
is
a
LOAEL,
the
LOAEL
can
be
used
for
the
RfD
calculation
with
the
inclusion
of
an
additional
uncertainty
factor.

UF
=
Uncertainty
factor
chosen
according
to
EPA/
NAS
guidelines
MF
=
Modifying
factor
Selection
of
the
uncertainty
factor
to
be
employed
in
calculation
of
the
RfD
is
based
on
professional
judgment,
while
considering
the
entire
database
of
toxicological
effects
for
the
chemical.
To
ensure
that
uncertainty
factors
are
selected
and
applied
in
a
consistent
manner,
the
Office
of
Water
(
OW)
employs
a
modification
to
the
guidelines
proposed
by
the
National
Academy
of
Sciences
(
NAS,
1977,
1980).
According
to
the
EPA
approach
(
U.
S.
EPA,
1993),

uncertainty
is
broken
down
into
its
components,
and
each
dimension
of
uncertainty
is
given
a
quantitative
rating.
The
total
uncertainty
factor
is
the
product
of
the
component
uncertainties.

The
individual
components
of
the
uncertainty
are
as
follows:

UF
H
A
factor
of
1,
3,
or
10­
fold
used
when
extrapolating
from
valid
data
in
studies
using
long­
term
exposure
to
average
healthy
humans.
This
factor
is
intended
to
account
for
the
variation
in
sensitivity
(
intraspecies
variation)
among
the
members
of
the
human
population.

UF
A
An
additional
factor
of
1,
3,
or
10
used
when
extrapolating
from
valid
results
of
long­
term
studies
on
experimental
animals
when
results
of
studies
of
human
exposure
are
not
available
or
are
inadequate.
This
factor
is
intended
to
account
for
the
uncertainty
involved
in
extrapolating
from
animal
data
to
humans
(
interspecies
variation).

UF
S
An
additional
factor
of
1,
3,
or
10
used
when
extrapolating
from
less­
thanchronic
results
on
experimental
animals
when
there
are
no
useful
long­
term
human
data.
This
factor
is
intended
to
account
for
the
uncertainty
involved
in
extrapolating
from
less­
than­
chronic
NOAELs
to
chronic
NOAELs.
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UF
L
An
additional
factor
of
1,
3,
or
10
used
when
deriving
an
RfD
from
a
LOAEL,
instead
of
a
NOAEL.
This
factor
is
intended
to
account
for
the
uncertainty
involved
in
extrapolating
from
LOAELs
to
NOAELs.

UF
D
An
additional
factor
of
1,
3­
or
10
used
when
deriving
an
RfD
from
an
"
incomplete"
database.
This
factor
is
meant
to
account
for
the
inability
of
any
single
type
of
study
to
consider
all
toxic
endpoints.
The
intermediate
factor
of
3
(
approximately
½
log
10
unit,
i.
e.,
the
square
root
of
10)
is
often
used
when
there
is
a
single
data
gap
exclusive
of
chronic
data.
It
is
often
designated
as
UF
D.

On
occasion,
EPA
also
uses
a
modifying
factor
in
the
determination
of
the
RfD.
A
modifying
factor
is
an
additional
uncertainty
factor
that
is
greater
than
zero
and
less
than
or
equal
to
10.
The
magnitude
of
the
MF
depends
upon
the
professional
assessment
of
scientific
uncertainties
of
the
study
and
database
not
explicitly
treated
above
(
e.
g.,
the
number
of
species
tested).
The
default
value
for
the
MF
is
1.

In
establishing
the
UF
or
MF,
it
is
recognized
that
professional
scientific
judgment
must
be
used.
The
total
product
of
the
uncertainty
factors
and
modifying
factor
should
not
exceed
3000.

If
the
assignment
of
uncertainty
results
in
a
UF/
MF
product
that
exceeds
3000,
then
the
database
does
not
support
development
of
an
RfD.
The
quantification
of
toxicological
effects
of
a
chemical
consists
of
separate
assessments
of
noncarcinogenic
and
carcinogenic
health
effects.

Unless
otherwise
specified,
chemicals
which
do
not
produce
carcinogenic
effects
are
believed
to
have
a
threshold
dose
below
which
no
adverse,
noncarcinogenic
health
effects
occur,
while
carcinogens
are
assumed
to
act
without
a
threshold.
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A.
1.1.
Drinking
Water
Equivalent
Level
The
drinking
water
equivalent
(
DWEL)
is
calculated
from
the
RfD.
The
DWEL
represents
a
drinking­
water­
specific
lifetime
exposure
at
which
adverse,
noncarcinogenic
health
effects
are
not
anticipated
to
occur.
The
DWEL
assumes
100%
exposure
from
drinking
water.

The
DWEL
provides
the
noncarcinogenic
health­
effects
basis
for
establishing
a
drinking­
water
standard.
For
ingestion
data,
the
DWEL
is
derived
as
follows:

DWEL
=
(
RfD)
×
BW
WI
where:

BW
=
70­
kg
adult
body
weight
WI
=
Drinking
water
intake
(
2
L/
day)

A.
1.2.
Health
Advisory
Values
In
addition
to
the
RfD
and
the
DWEL,
EPA
calculates
Health
Advisory
(
HA)
values
for
noncancer
effects.
HAs
are
determined
for
lifetime
exposures
as
well
as
for
exposures
of
shorter
duration
(
1­
day,
10­
day,
and
longer­
term).
The
shorter
duration
HA
values
are
used
as
informal
guidance
to
municipalities
and
other
organizations
when
emergency
spills
or
contamination
situations
occur.
The
lifetime
HA
becomes
the
MCLG
for
a
chemical
that
is
not
a
carcinogen.

The
shorter­
term
HAs
are
calculated
using
an
equation
similar
to
the
ones
for
RfD
and
DWEL;
however,
the
NOAELs
or
LOAELs
are
derived
from
acute
or
subchronic
studies
and
identify
a
sensitive
noncarcinogenic
endpoint
of
toxicity.
The
HAs
are
derived
as
follows:
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HA
=
NOAEL
or
LOAEL
×
BW
UF
×
WI
where:

NOAEL
or
LOAEL
=
No­
or
lowest­
observed­
adverse­
effect­
level
in
mg/
kg
bw/
day
BW
=
Assumed
body
weight
of
a
child
(
10
kg)
or
an
adult
(
70
kg)

UF
=
Uncertainty
factor,
in
accordance
with
EPA
or
NAS/
OW
guidelines
WI
=
Assumed
daily
water
consumption
of
a
child
(
1
L/
day)
or
an
adult
(
2
L/
day)

Using
the
above
equation,
the
following
drinking
water
HAs
are
developed
for
noncarcinogenic
effects:

°
1­
day
HA
for
a
10­
kg
child
ingesting
1
L
water
per
day.

°
10­
day
HA
for
a
10­
kg
child
ingesting
1
L
water
per
day.

°
Longer­
term
HA
for
a
10­
kg
child
ingesting
1
L
water
per
day.

°
Longer­
term
HA
for
a
70­
kg
adult
ingesting
2
L
water
per
day.

Each
of
these
shorter­
term
HA
values
assumes
that
the
total
exposure
to
the
contaminant
comes
from
drinking
water.

The
lifetime
HA
is
calculated
from
the
DWEL
and
takes
into
account
exposure
from
sources
other
than
drinking
water.
It
is
calculated
using
the
following
equation:
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Lifetime
HA
=
DWEL
×
RSC
where:

DWEL
=
Drinking
water
equivalent
level
RSC
=
Relative
source
contribution.
The
fraction
of
the
total
exposure
allocated
to
drinking
water
following
EPA
guidance
(
U.
S.
EPA,
2000b).

A.
2
Quantification
of
Carcinogenic
Effects
Under
the
1986
guidelines,
the
EPA
categorizes
the
carcinogenic
potential
of
a
chemical
based
on
the
overall
weight­
of­
evidence
according
to
the
following
scheme:

°
Group
A:
Human
Carcinogen.
Sufficient
evidence
exists
from
epidemiology
studies
to
support
a
causal
association
between
exposure
to
the
chemical
and
human
cancer.

°
Group
B:
Probable
Human
Carcinogen.
Sufficient
evidence
of
carcinogenicity
in
animals
with
limited
(
Group
B1)
or
inadequate
(
Group
B2)
evidence
in
humans.

°
Group
C:
Possible
Human
Carcinogen.
Limited
evidence
of
carcinogenicity
in
animals
in
the
absence
of
human
data.

°
Group
D:
Not
classified
as
to
Human
Carcinogenicity.
Inadequate
human
and
animal
evidence
of
carcinogenicity
or
for
which
no
data
are
available.

°
Group
E:
Evidence
of
Noncarcinogenicity
for
Humans.
No
evidence
of
carcinogenicity
in
at
least
two
adequate
animal
tests
in
different
species
or
in
both
adequate
epidemiologic
and
animal
studies.
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If
toxicological
evidence
leads
to
the
classification
of
the
contaminant
as
a
genotoxic,

probable
or
possible
human
carcinogen,
mathematical
models
are
used
to
calculate
the
estimated
excess
cancer
risk
associated
with
ingestion
of
the
contaminant
in
drinking
water.
The
data
used
in
these
estimates
usually
come
from
lifetime­
exposure
studies
in
animals.
In
order
to
predict
the
risk
for
humans
from
animal
data,
animal
doses
must
be
converted
to
equivalent
human
doses.

This
conversion
includes
correction
for
noncontinuous
exposure,
less­
than­
lifetime
studies
and
differences
in
size.
It
is
assumed
that
the
average
adult
human­
body
weight
is
70
kg
and
that
the
average
water
consumption
of
an
adult
human
is
two
liters
of
water
per
day.

For
contaminants
with
a
carcinogenic
potential,
chemical
levels
are
correlated
with
a
carcinogenic­
risk
estimate
by
employing
a
cancer
potency
(
unit
risk)
value
together
with
the
assumption
for
lifetime
exposure
via
ingestion
of
water.
Under
the
1986
Carcinogen
Risk
Assessment
Guidelines,
the
cancer
unit
risk
is
usually
derived
from
a
linearized
multistage
model
with
a
95%
upper
confidence
limit
providing
a
low­
dose
estimate;
that
is,
the
true
risk
to
humans,

while
not
identifiable,
is
not
likely
to
exceed
the
upper­
limit
estimate
and,
in
fact,
may
be
lower.

Excess
cancer­
risk
estimates
may
also
be
calculated
using
other
models
such
as
the
one­
hit,

Weibull,
logit
and
probit
models.
There
is
little
basis
in
the
current
understanding
of
the
biological
mechanisms
involved
in
cancer
to
suggest
that
any
one
of
these
models
is
able
to
predict
risk
more
accurately
than
any
of
the
others.
Because
each
model
is
based
upon
differing
assumptions,
the
estimates
that
are
derived
for
each
model
can
differ
by
several
orders
of
magnitude.

The
scientific
data
base
used
to
calculate
and
support
the
setting
of
cancer­
risk
rates
has
an
inherent
uncertainty
due
to
the
systematic
and
random
errors
in
scientific
measurement.
In
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most
cases,
only
studies
using
experimental
animals
have
been
performed.
Thus,
there
is
uncertainty
when
the
data
are
extrapolated
to
humans.
When
developing
cancer­
risk
rates,
several
other
areas
of
uncertainty
exist,
such
as
the
incomplete
knowledge
concerning
the
health
effects
of
contaminants
in
drinking
water,
the
impact
of
the
experimental
animal's
age,
sex
and
species,
the
nature
of
the
target
organ
system(
s)
examined
and
the
actual
rate
of
exposure
of
the
internal
targets
in
experimental
animals
or
humans.
Dose­
response
data
usually
are
available
only
for
high
levels
of
exposure,
not
for
the
lower
levels
of
exposure
at
which
a
standard
may
be
set.
When
there
is
exposure
to
more
than
one
contaminant,
additional
uncertainty
results
from
a
lack
of
information
about
possible
synergistic
or
antagonistic
effects.

The
quantification
of
toxicological
effects
of
a
chemical
consists
of
separate
assessments
of
noncarcinogenic
and
carcinogenic
health
effects.
Chemicals
that
do
not
produce
carcinogenic
effects
are
believed
to
have
a
threshold
dose
below
which
no
adverse,
noncarcinogenic
health
effects
occur,
while
carcinogens
are
assumed
to
act
without
a
threshold.

B.
Noncarcinogenic
Effects
Analysis
of
the
dose­
response
data
for
noncarcinogenic
effects
for
each
of
the
four
HANs
and
the
derivation
of
Health
Advisories
is
described
below
and
summarized
in
Tables
VIII­
1
through
VIII­
7.

B.
1
BCAN
The
oral
toxicity
data
for
BCAN
are
summarized
in
Table
VIII­
1.
No
systemic
toxicity
studies
of
BCAN
are
available.
The
most
comprehensive
studies
of
BCAN
toxicity
were
two
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developmental
studies
of
BCAN.
Smith
et
al.
(
1987)
reported
that
the
single
dose
tested
of
55
mg/
kg/
day
of
BCAN,
administered
by
gavage
to
pregnant
rats
on
days
7
to
21
of
gestation,

resulted
in
decreased
maternal
weight
gain
and
reduced
pup
birth
weights.
In
a
dose­
response
study
(
Christ
et
al.,
1995),
sperm­
positive
female
rats
were
administered
BCAN
by
gavage
in
tricaprylin
on
gestation
days
6
to
18
at
doses
of
0,
5,
25,
45,
and
65
mg/
kg/
day.
Treatment
with
BCAN
in
tricaprylin
resulted
in
both
maternal
and
embryotoxicity.
The
LOAEL
for
developmental
effects
was
5
mg/
kg/
day
compared
to
tricaprylin
treated
controls.
No
maternal
effects
were
observed
at
this
dose.
However,
use
of
this
study
for
dose­
response
assessment
is
not
appropriate,
because
it
may
not
accurately
reflect
the
toxicity
of
BCAN
in
drinking
water.

Tricaprylin
vehicle
alone
produced
embryotoxicity
in
this
study,
and
later
work
by
this
laboratory
(
Christ
et
al.,
1996)
suggests
that
tricaprylin
may
act
synergistically
with
TCAN
to
enhance
developmental
toxicity.
In
the
absence
of
a
more
complete
data
base,
the
data
are
inadequate
for
derivation
of
any
Health
Advisory
values
for
BCAN.
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Draft
VIII­
10
Table
VIII­
1
Summary
of
Oral
Studies
of
BCAN
Toxicity
Reference
Species/
Strain
Route/
Dose
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

Meier
et
al.
(
1985)
Mouse­

B6C3F1
Gavage
in
water
0,
12.5,
25,
or
50
mg/
kg/
day
5
Days
Sperm
head
abnormalities
50
(
Freestanding
NOAEL)
NDa
Smith
et
al.
(
1987)
Rat­

Long­
Evans
Hooded
Gavage
in
tricaprylinb
55
mg/
kg/
day
Days
7
to
21
of
gestation
Maternal
weight,
reproductive
success,
pup
viability
and
growth
Maternal:
ND
Developmental:
ND
Maternal:
ND
(
Nonsignificant
decrease
maternal
weight
gain)

Development:
55
(
Decreased
birth
weight,
decreased
postnatal
weight
gain)

Christ
et
al.
(
1995)
Rat­

Long­
Evans
Gavage
in
tricaprylinb
0,
5,
25,
45,
65
mg/
kg/
day
Days
6
to
18
of
gestation
Maternal
body
and
organ
weight,
reproductive
success,
pup
viability
and
growth,
malformations
Maternal:
45
Developmental:
ND
Maternal:
65
(
FEL
for
maternal
death;
decrease
maternal
weight
gain)

Development:
5
(
Decreased
crownrump
length,
increased
cardiovascular
malformations)

a.
ND
=
not
determined.
b.
Due
to
the
demonstrated
ability
of
tricaprylin
to
enhance
developmental
toxicity
of
TCAN,
this
study
was
not
considered
in
derivation
of
the
Health
Advisories.

B.
2
DBAN
B.
2.1
One­
Day
Health
Advisory
for
DBAN
The
oral
toxicity
data
for
DBAN
are
summarized
in
Table
VIII­
2.
Two
studies
(
Hayes
et
al.,
1986;
Eastman
Kodak
Co.,
1992)
identify
acute
oral
LD
50
values
for
DBAN
of
50
to
361
mg/
kg.
Clinical
signs
observed
include
convulsions,
ataxia,
depressed
respiration
and
activity,
and
coma.
However,
LD
50
studies
are
not
suitable
for
the
development
of
one­
day
health
advisories
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and
no
other
suitable
acute
studies
were
located
for
DBAN.
In
the
absence
of
such
data,
the
Ten­
day
HA
value
is
recommended
as
a
conservative
estimate
of
an
appropriate
One­
day
HA
value.

B.
2.2
Ten­
Day
Health
Advisory
for
DBAN
Two
reports
of
adequate
general
toxicity
studies
(
NTP,
2002,
which
tested
both
mice
and
rats,
and
Hayes
et
al.,
1986,
which
tested
rats
only)
of
a
suitable
duration
for
the
Ten­
day
HA
were
located.
NTP
(
2002)
conducted
a
14­
day
drinking
water
toxicity
study
in
B6C3F1
mice
and
F344
rats.
Concentration­
related
decreases
in
water
consumption
were
noted
in
males
and
females
of
both
species,
but
this
effect
was
not
considered
to
be
toxicologically
significant.
The
only
toxicologically­
significant
effects
in
either
species
were
observed
in
male
rats
at
the
high
dose
of
18
mg/
kg/
day,
and
included
decreased
body
weight,
decreased
testes
weight
and
testes
atrophy.
The
NOAEL
for
this
study
was
12
mg/
kg/
day
with
a
LOAEL
of
18
mg/
kg/
day.
BMD
modeling
was
not
performed
for
this
study,
since
the
full
NTP
study
reports
were
not
available
at
the
time
of
preparation
of
this
document.
However,
preliminary
modeling
based
on
the
body
weight
gain
data
provided
in
the
study
summaries
suggests
that
the
BMDL
would
not
differ
significantly
from
the
study
NOAEL.

Hayes
et
al.
(
1986)
conducted
a
14­
day
study
of
DBAN
toxicity
in
rats.
No
significant
effects
on
serum
chemistry,
hematological
or
urinary
parameters
or
remarkable
findings
at
necropsy
were
observed.
The
only
organ
weight
change
that
showed
a
clear
dose
dependence
was
relative
liver
weight
in
females,
which
was
increased
12%
over
controls
(
p

0.05)
at
23
mg/
kg/
day,
and
was
increased
by
as
much
as
22%
above
controls
at
90
mg/
kg/
day.
The
increase
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12
in
relative
liver
weight
was
statistically
significant
only
at
the
high
dose.
In
the
absence
of
histopathology
data
or
clinical
chemistry
findings,
however,
it
is
unclear
if
this
is
an
adverse
response.
In
males,
body
weight
was
decreased
at
the
45
mg/
kg/
day
dose
level,
but
not
at
the
23
mg/
kg/
day
dose
level.
Therefore,
decreased
body
weight
in
males
is
considered
the
critical
effect
and
23
mg/
kg/
day
is
considered
the
NOAEL.
BMD
modeling
was
conducted
to
identify
alternative
critical
effect
levels
for
this
study.
A
BMDL
of
16
mg/
kg/
day
for
decreased
body
weight
in
males
was
selected
as
the
most
appropriate
modeling
result
for
this
endpoint
(
see
Appendix
A).

As
part
of
a
reproductive
toxicity
study,
R.
O.
W.
Sciences
(
1997)
conducted
range­
finding
studies
of
DBAN
that
evaluated
its
short­
term
effects.
Among
rats
exposed
to
drinking
water
concentrations
of
DBAN
up
to
200
ppm
for
2
weeks,
the
only
consistent
effect
was
a
decrease
in
water
consumption
at
the
high
concentration.
The
absence
of
clinical
signs
of
toxicity
or
body
weight
changes
indicates
that
the
highest
concentration
tested
of
200
ppm
in
the
second
rangefinding
study
(
equivalent
to
doses
of
13.2
mg/
kg/
day
in
males;
17.9
mg/
kg/
day
in
females)
is
a
study
NOAEL.
These
same
rats
had
previously
been
exposed
for
4
days
to
higher
concentrations
that
caused
significant
decreases
in
body
weight,
and
decreased
food
and
water
consumption;
the
rats
were
allowed
to
recover
to
control
body
weights
before
being
exposed
to
the
lower
concentrations
in
the
second
range­
finding
study.
In
the
main
reproductive
and
developmental
study
the
exposure
duration
was
30
days
for
males
and
35
days
for
females
­
longer
than
is
suitable
for
a
Ten­
day
HA.
In
addition,
there
were
no
adverse
effects
observed
at
the
highest
dose
tested,
8.2
mg/
kg/
day
in
males
and
10.8
mg/
kg/
day
in
females.
Smith
et
al.
(
1987)
reported
on
the
developmental
toxicity
of
DBAN.
However,
use
of
this
study
for
dose­
response
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assessment
is
not
appropriate,
because
it
may
not
accurately
reflect
the
toxicity
of
DBAN
in
drinking
water,
due
to
interactions
between
tricaprylin
vehicle
and
HANs.

Decreased
body
weight,
decreased
testes
weight,
and
testes
atrophy
in
male
F344
rats
reported
in
NTP
(
2002)
at
doses
greater
than12
mg/
kg/
day
is
considered
the
most
appropriate
basis
for
deriving
the
Ten­
day
HA
for
DBAN.
Decreased
body
weight
was
clearly
an
appropriate
endpoint
to
serve
as
the
critical
effect
for
deriving
the
Ten­
day
HA.
Although
the
reported
effects
of
DBAN
on
the
testes
were
considered
to
be
adverse,
their
appropriateness
to
serve
as
the
basis
for
the
HA
was
not
clear.
The
Ten­
day
HA
is
based
on
water
consumption
by
children
directly.

Therefore,
male
reproductive
effects
are
not
an
appropriate
endpoint
for
this
HA
value
unless
the
observed
effects
are
likely
to
persist
to
a
reproductive
age.
The
data
were
not
adequate
to
make
this
determination,
since
none
of
the
shorter­
term
studies
tracked
the
recovery
of
this
endpoint
after
cessation
of
exposure.
It
is
noteworthy
that
no
effects
on
the
testes
were
observed
in
the
13­

week
NTP
(
2002)
study
in
the
same
strain
of
rats,
suggesting
that
the
testes
effects
might
be
transient.
However,
the
highest
dose
in
the
subchronic
study
NTP
(
2002)
study
was
11.3
mg/
kg/
day,
which
is
essentially
the
same
as
the
NOAEL
for
testes
effects
in
the
14­
day
NTP
study.
Therefore,
comparison
across
the
14­
day
and
13­
week
NTP
studies
cannot
answer
whether
the
testes
effects
are
likely
to
be
persistent.
Hayes
et
al.
(
1986)
also
evaluated
testes
weight
in
CD
rats
after
14­
day
and
subchronic
gavage
dosing
with
DBAN.
This
study
did
not
identify
decreased
testes
weight
at
either
time
point.
This
could
be
due
to
rat
strain
differences
in
sensitivity
to
male
reproductive
tract
toxicity,
or
reflect
differences
in
the
route
of
DBAN
administration.
DBAN
administered
in
drinking
water
caused
a
decrease
in
water
consumption
in
the
NTP
(
2002)
study.
If
the
effects
on
the
testes
were
secondary
to
decreased
water
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consumption,
then
they
would
not
have
been
observed
in
the
Hayes
et
al.
(
1986)
study,
which
used
gavage
dosing.
DBAN
did
not
affect
male
reproductive
parameters
in
Sprague­
Dawley
rats
in
a
reproductive
and
developmental
screening
study
(
R.
O.
W.
Sciences,
1997)
in
which
males
were
exposed
for
30
days
to
DBAN
in
drinking
water.
However,
in
this
study,
the
highest
dose
tested
was
8.2
mg/
kg/
day
(
the
drinking
water
concentration
was
150
mg/
L),
which
was
below
the
NOAEL
of
12
mg/
kg/
day
for
F344
rats
in
the
14­
day
NTP
study.
None
of
the
available
studies
allow
for
a
determination
of
the
degree
to
which
the
testes
effects
are
likely
to
persist.
Therefore,

since
the
potential
of
the
observed
testicular
effects
to
persist
to
a
reproductive
age
cannot
be
excluded,
they
are
considered
to
be
appropriate
co­
critical
effects
for
derivation
of
the
Ten­
day
HA.

Based
on
the
decreased
body
weight,
decreased
testes
weight,
and
testes
atrophy
in
male
rats
reported
in
NTP
(
2002)
at
doses
greater
than
12
mg/
kg/
day,
the
Ten­
day
HA
for
DBAN
may
be
calculated
as
shown
below
and
summarized
in
Table
VIII­
3.
An
uncertainty
factor
of
10
is
used
to
account
for
extrapolation
from
a
NOAEL
in
an
animal
study
and
an
uncertainty
factor
of
10
is
used
to
account
for
inter­
individual
variability
in
human
sensitivity.
The
composite
uncertainty
factor
used
is
100.

(
12
mg/
kg/
day)
(
10
kg)
Ten­
day
HA
=
=
1.2
mg/
L
(
rounded
to
1
mg/
L)
(
100)
(
1
L/
day)

where:

12
mg/
kg/
day
=
NOAEL,
based
on
decreased
body
weight,
decreased
testes
weight,
and
testes
atrophy
in
male
rats
exposed
to
a
LOAEL
of
18
mg/
kg/
day
DBAN
in
drinking
water
for
14
days
(
NTP,
2002).

10
kg
=
assumed
body
weight
of
a
child.
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100
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
NOAEL
in
animals,
and
inter­
individual
variability
in
humans.

1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

B.
2.3
Longer­
Term
Health
Advisory
for
DBAN
Only
two
reports
of
suitable
studies
for
deriving
a
longer­
term
HA
were
located.
NTP
(
2002)
evaluated
the
subchronic
toxicity
of
DBAN
in
B6C3F1
mice
and
F344
rats
exposed
to
DBAN
in
their
drinking
water.
In
both
species
the
only
effects
were
decreased
water
consumption
and
slight
decreases
in
body
weight.
The
highest
dose
tested
was
the
study
NOAEL
of
17.9
mg/
kg/
day
for
both
male
and
female
mice,
12.6
mg/
kg/
day
for
female
rats,
and
11.3
mg/
kg/
day
for
male
rats.
No
NOAEL
was
identified.

Hayes
et
al.
(
1986)
conducted
a
90­
day
gavage
study
of
DBAN
toxicity
in
CD
rats.
At
the
high
dose
of
45
mg/
kg/
day,
males,
but
not
females,
had
decreased
body
weight.
The
next
lower
dose
of
23
mg/
kg/
day
was
the
NOAEL
for
decreased
body
weight.
The
only
other
noteworthy
effects
observed
in
the
study
were
significantly
increased
ALP
in
females
at
45
mg/
kg/
day
and
a
significant
increase
in
relative
(
but
not
absolute)
liver
weight
in
males
at
45
mg/
kg/
day.
With
the
exception
of
elevated
ALP
levels
at
the
high
dose,
there
were
no
significant
treatment­
related
effects
on
serum
chemistry,
hematological,
or
urinary
parameters
or
remarkable
findings
at
necropsy
at
any
dose
level.
The
observed
liver
weight
changes
were
not
judged
as
adverse
since
no
clinical
chemistry
signs
of
liver
toxicity
were
observed
in
males.
Females
had
an
increase
in
ALP
at
the
high
dose,
but
did
not
have
a
corresponding
increase
in
liver
weight.
No
histopathology
examination
was
performed
to
clarify
if
the
liver
weight
changes
in
males
was
Drinking
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adverse
or
adaptive.
Based
on
these
considerations,
decreased
body
weight
in
males
was
selected
as
the
critical
effect
for
this
study.
The
NOAEL
was
23
mg/
kg/
day
and
the
LOAEL
was
45
mg/
kg/
day.
BMD
modeling
was
conducted
for
decreased
body
weight
in
males
to
identify
alternative
critical
effect
levels
for
this
study.
A
BMDL
of
20
mg/
kg/
day
was
selected
as
the
most
appropriate
modeling
result
for
this
endpoint
(
see
Appendix
A).

Both
the
NTP
(
2002)
subchronic
drinking
water
study
in
male
rats
and
the
subchronic
gavage
study
by
Hayes
et
al.
(
1986)
were
considered
in
the
selection
of
the
critical
study
for
derivation
of
the
Longer­
term
HA.
The
NOAEL
of
11.3
mg/
kg/
day
for
male
rats
in
the
13­
week
NTP
study
was
selected
as
the
most
appropriate
basis
for
derivation
of
the
Longer­
term
HA.
This
value
was
judged
to
be
more
appropriate
for
deriving
the
HA
than
the
NOAEL
of
23
mg/
kg/
day
for
decreased
body
weight
observed
in
male
rats
reported
in
Hayes
et
al.
(
1986)
for
several
reasons.
First,
in
the
NTP
study
DBAN
was
administered
in
drinking
water,
a
dose
route
more
relevant
to
environmental
exposure
than
the
corn­
oil
gavage
dosing
employed
by
Hayes
et
al.

(
1986).
Second,
although
the
NTP
13­
week
study
did
not
identify
a
LOAEL,
the
NOAELs
for
decreased
body
weight
were
the
same
for
the
14­
day
and
13­
week
NTP
studies,
and
the
LOAEL
was
18
mg/
kg/
day
in
the
14­
day
study.
Since
slight
body
weight
decreases
were
also
observed
in
the
13­
week
study
at
11.3
mg/
kg/
day,
this
suggests
that
the
LOAEL
for
the
13­
week
study
might
approximate
the
LOAEL
of
18
mg/
kg/
day
for
the
14­
day
study,
which
is
significantly
lower
than
the
LOAEL
of
45
mg/
kg/
day
reported
in
Hayes
et
al.
(
1986).
This
argues
that
the
NOAEL/
LOAEL
boundary
would
be
lower
in
the
NTP
(
2002)
study
than
in
Hayes
et
al.
(
1986).

Third,
since
the
NTP
(
2002)
and
Hayes
et
al.
(
1986)
studies
are
not
directly
comparable,
due
to
differences
in
the
methods
of
dose
administration
and
rats
strains
employed,
and
both
studies
were
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
17
of
adequate
quality
to
derive
the
Longer­
term
HA,
selection
of
the
lower
study
NOAEL
would
be
most
appropriate,
even
in
the
absence
of
a
LOAEL.

The
Longer­
term
HA
value
for
a
10­
kg
child
for
DBAN
may
be
calculated
as
shown
below
and
summarized
in
Table
VIII­
3.
Derivation
of
the
health
advisories
is
shown
using
the
study
NOAEL
of
11.3
mg/
kg/
day
from
the
NTP
(
2002)
13­
week
study
as
the
point
of
departure.

An
uncertainty
factor
of
10
is
used
to
account
for
extrapolation
from
an
animal
study
and
an
uncertainty
factor
of
10
is
used
to
account
for
inter­
individual
variability
in
human
sensitivity,
in
the
absence
of
sufficient
data
to
depart
from
these
defaults.
An
additional
uncertainty
factor
of
3
is
used
to
account
for
database
insufficiencies.
This
factor
is
selected
since
none
of
the
available
reproductive
or
developmental
studies
were
adequate
to
use
in
the
quantitative
dose­
response
assessment.
The
data
gap
may
be
particularly
relevant
since
cyanide,
a
metabolite
of
DBAN,

induces
male
reproductive
system
toxicity
(
U.
S.
EPA,
2002c),
and
due
to
uncertainty
regarding
the
significance
of
the
testes
effects
observed
in
the
14­
day
NTP
(
2002)
study
for
DBAN.
The
reproductive
and
developmental
toxicity
study
by
R.
O.
W.
Sciences
(
1997)
was
limited
by
the
fact
that
this
was
a
screening
study
that
was
not
designed
to
evaluate
the
full
spectrum
of
endpoints
of
interest.
The
developmental
toxicity
study
by
Smith
et
al.
(
1987)
is
of
limited
use,
because
it
was
a
single­
dose
study,
because
an
insufficient
array
of
endpoints
was
evaluated,
and
because
the
observed
toxicity
was
confounded
by
the
use
of
tricaprylin
as
the
solvent
vehicle.
A
full
factor
of
10
was
not
used
for
the
database
uncertainty
factor
since
the
systemic
toxicity
of
DBAN
has
been
tested
in
two
species
in
subchronic
studies
(
NTP,
2002;
Hayes
et
al.,
1986).
The
composite
uncertainty
factor
used
is
300.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
18
Derivation
of
the
Longer­
term
HA
based
on
the
study
NOAEL
(
11.3
mg/
kg/
day)
(
10
kg)
Longer­
Term
HA
=
=
0.38
mg/
L
(
rounded
to
0.4
mg/
L)
(
300)
(
1
L/
day)

where:

11.3
mg/
kg/
day
=
NOAEL
in
male
F344
rats
exposed
to
DBAN
in
drinking
water
for
13­
weeks
(
NTP,
2002).

10
kg
=
assumed
body
weight
of
a
child.

300
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
NOAEL
in
animals,
inter­
individual
variability
in
humans,
and
insufficiencies
in
the
database.

1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

The
Longer­
term
HA
for
a
70­
kg
adult
consuming
2
L/
day
of
water
is
calculated
as
follows:

(
11.3
mg/
kg/
day)
(
70
kg)
Longer­
Term
HA
=
=
1.3
mg/
L
(
rounded
to
1
mg/
L)
(
300)(
2
L/
day)

where:

11.3
mg/
kg/
day
=
NOAEL
in
male
F344
rats
exposed
to
DBAN
in
drinking
water
for
13­
weeks
(
NTP,
2002).

70
kg
=
assumed
body
weight
of
an
adult.

300
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
NOAEL
in
animals,
inter­
individual
variability
in
humans,
and
insufficiencies
in
the
database.

2
L/
day
=
assumed
daily
water
consumption
by
a
70­
kg
adult.

B.
2.4
Lifetime
Health
Advisory
for
DBAN
No
chronic
studies
of
DBAN
toxicity
were
located,
although
DBAN
is
currently
under
test
for
chronic
toxicity
in
mice
and
rats
(
NTP,
2002).
In
the
absence
of
such
data,
the
available
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
19
subchronic
studies
(
NTP,
2002;
Hayes
et
al.,
1986)
may
be
used
to
derive
the
Lifetime
HA.
As
described
for
the
Longer­
term
HA,
the
NOAEL
of
11.3
mg/
kg/
day
for
male
rats
identified
in
the
NTP
(
2002)
study
is
the
most
appropriate
basis
for
deriving
the
RfD.
The
derivation
of
the
RfD
is
shown
below
and
summarized
in
Table
VIII­
3.
An
uncertainty
factor
of
10
is
used
to
account
for
extrapolation
from
an
animal
study
and
an
uncertainty
factor
of
10
is
used
to
account
for
interindividual
variability
in
human
sensitivity,
in
the
absence
of
sufficient
data
to
depart
from
these
defaults.
An
uncertainty
factor
of
3,
instead
of
the
default
value
of
10,
was
chosen
to
account
for
less­
than­
lifetime
exposure,
based
on
the
absence
of
progression
of
toxicological
effects
(
or
even
regression)
from
14
days
to
90
days
(
NTP,
2002;
Hayes
et
al.,
1986).
An
uncertainty
factor
of
3
is
used
to
account
for
insufficiencies
in
the
database.
This
factor
was
chosen
to
replace
the
default
factor
of
10
because
the
subchronic
toxicity
of
DBAN
has
been
evaluated
in
two
species
(
NTP,

2002;
Hayes
et
al.,
1986).
Furthermore,
decreased
body
weight
was
the
identified
as
the
most
sensitive
effect
in
both
studies,
even
though
the
NTP
study
included
a
thorough
examination
of
tissue
histopathology,
hematology,
and
clinical
chemistry.
These
results
suggest
that
no
new
systemic
target
organs
for
DBAN
are
likely
to
be
identified.
However,
none
of
the
available
reproductive
or
developmental
studies
were
adequate
to
use
in
the
quantitative
dose­
response
assessment.
The
data
gap
may
be
particularly
relevant
since
cyanide,
a
metabolite
of
DBAN,

induces
male
reproductive
system
toxicity
(
U.
S.
EPA,
2002c),
and
due
to
uncertainty
regarding
the
significance
of
the
testes
effects
observed
in
the
14­
day
NTP
(
2002)
study
for
DBAN.
The
reproductive
and
developmental
toxicity
study
by
R.
O.
W.
Sciences
(
1997)
was
limited
by
the
fact
that
this
was
a
screening
study
that
was
not
designed
to
evaluate
the
full
spectrum
of
endpoints
of
interest.
The
developmental
toxicity
study
by
Smith
et
al.
(
1987)
is
of
limited
use,
because
it
was
a
single­
dose
study,
because
an
insufficient
array
of
endpoints
was
evaluated,
and
because
the
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
20
observed
toxicity
was
confounded
by
the
use
of
tricaprylin
as
the
solvent
vehicle.
Therefore,

based
on
default
factors
of
10
each
for
interspecies
extrapolation
and
inter­
individual
variability,

and
partial
factors
of
3
each
for
subchronic
to
chronic
extrapolation
and
for
database
insufficiencies
(
lack
of
adequate
developmental
and
reproductive
toxicity
studies),
the
composite
uncertainty
factor
used
is
1000.

Derivation
of
the
Lifetime
HA
based
on
the
study
NOAEL.

Step
1:
Determination
of
the
Reference
Dose
(
RfD)
for
DBAN
(
11.3
mg/
kg/
day)
RfD
=
=
0.011
mg/
kg/
day
(
rounded
to
0.01
mg/
kg/
day)
(
1000)

where:

11.3
mg/
kg/
day
=
NOAEL
in
male
F344
rats
exposed
to
DBAN
in
drinking
water
for
13­
weeks
(
NTP,
2002).

1000
=
composite
uncertainty
factor
chosen
to
account
for
extrapolation
from
a
NOAEL
in
animals,
inter­
individual
variability
in
humans,
less­
than­
lifetime
exposure,
and
insufficiencies
in
the
database.

Step
2:
Determination
of
the
Drinking
Water
Equivalent
Level
(
DWEL)
for
DBAN
(
0.011
mg/
kg/
day)
(
70
kg)
DWEL
=
=
0.39
mg/
L
(
rounded
to
0.4
mg/
L)
(
2
L/
day)

where:

0.011
mg/
kg/
day
=
RfD
(
before
rounding)

70
kg
=
assumed
body
weight
of
an
adult.

2
L/
day
=
assumed
daily
water
consumption
by
a
70­
kg
adult.

Step
3:
Determination
of
the
Lifetime
Health
Advisory
for
DBAN
Lifetime
HA
=
(
0.39
mg/
L)
(
20%)
=
0.078
mg/
L
(
rounded
to
0.08
mg/
L)
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
21
where:
0.39
mg/
L
=
DWEL
20%
=
assumed
relative
source
contribution
from
water
Table
VIII­
2
Summary
of
Oral
Studies
of
DBAN
Toxicity
Reference
Species/
Strain
Route/
Dose
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

Hayes
et
al.
(
1986)
Mouse­

B6C3F1
Gavage
in
corn
oil
25
­
3,200
mg/
kg/
day
Acute
Lethality
NDa
LD50
=
289
(
M)
303
(
F)

Rat­

CD
Gavage
in
corn
oil
25
­
1,600
mg/
kg/
day
Acute
Lethality
ND
LD50
=
245
(
M)
361
(
F)

Eastman
Kodak
Co.
(
1992)
Mouse
Not
specified
Gavage
25
­
1,600
mg/
kg/
day
Acute
Lethality
ND
LD50
=
50
Rat
Not
specified
Gavage
25
­
3200
mg/
kg/
day
Acute
Lethality
ND
LD50
=
50
­
100
Meier
et
al.
(
1985)
Mouse­

B6C3F1
Gavage
in
water
0,
12.5,
25,
or
50
mg/
kg/
day
5
Days
Sperm
head
abnormalities
50
(
Freestanding
NOAEL)
ND
R.
O.
W
Sciences
(
1997)
Rat­

Sprague­
Dawley
Drinking
Water
0,
0.7,
2.2,
5.8,
13.2
mg/
kg/
day
(
males)

0,
0.8,
2.4,
6.8,
17.9
mg/
kg/
day
(
females)
14
Days
Clinical
signs,
body
weight,
food
consumption
13.2
(
m);
17.9
(
f)
(
Freestanding
NOAEL)
ND
Hayes
et
al.
(
1986)
Rat­

CD
Gavage
in
corn
oil
0,
23,
45,
90,
180
mg/
kg/
day
14
Days
Body
weight,
organ
weight,
serum
chemistry,
hematology,
urinalysis,
gross
necropsy
23
45
(
Decreased
body
weight
in
males)
Drinking
Water
Criteria
Document
for
Haloacetonitriles
Reference
Species/
Strain
Route/
Dose
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

EPA/
OW/
OST/
HECD
Final
Draft
VIII­
22
Rat­

CD
Gavage
in
corn
oil
0,
6,
23,
45
mg/
kg/
day
90
Days
Body
weight,
organ
weight,
serum
chemistry,
hematology,
urinalysis,
gross
necropsy
23
45
(
Decreased
body
weight
in
males)

NTP
(
2002)
Mice­

B6C3F1
Drinking
Water
0,
2.1,
4.3,
8.2,
14.7,
21.4
mg/
kg/
day
(
Males)

0,
2.0,
3.3,
10.0,
13.9,
21.6
mg/
kg/
day
(
Females)
14
Days
Clinical
signs,
body
weight,
water
consumption,
organ
weight
and
pathology,
liver
GST
activity
21
(
Freestanding
NOAEL)

Rat­

Fischer­
344
Drinking
Water
0,
2,
3,
7,
12,
18
mg/
kg/
day
(
Males)

0,
2,
4,
7,
12,
19
mg/
kg/
day
(
Females)
14
Days
Clinical
signs,
body
weight,
water
consumption,
organ
weight
and
pathology,
liver
GST
activity
12
(
m)
18
(
Decreased
body
weight,
decreased
testes
weight
and
pathology
in
males)

Mice­

B6C3F1
Drinking
Water
0,
1.6,
3.2,
5.6,
10.7,
17.9
mg/
kg/
day
(
Males)

0,
1.6,
3,
6.1,
11.1,
17.9
mg/
kg/
day
(
Females)
13
Weeks
Clinical
signs,
body
weight,
water
consumption,
organ
weight
and
pathology,
hematology
and
clinical
chemistry
17.9
(
m)
(
Freestanding
NOAEL)

Rat­

Fischer­
344
Drinking
Water
0,
0.9,
1.8,
3.3,
6.2,
11.3
mg/
kg/
day
(
Males)

0,
1,
1.9,
3.8,
6.8,
12.6
mg/
kg/
day
(
Females)
13
Weeks
Clinical
signs,
body
weight,
water
consumption,
organ
weight
and
pathology,
hematology
and
clinical
chemistry
11.3
(
m);
12.6
(
f)
(
Freestanding
NOAEL)
Drinking
Water
Criteria
Document
for
Haloacetonitriles
Reference
Species/
Strain
Route/
Dose
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

EPA/
OW/
OST/
HECD
Final
Draft
VIII­
23
R.
O.
W
Sciences
(
1997)
Rat­

Sprague­
Dawley
Drinking
Water
0,
1.4,
3.3,
8.2
mg/
kg/
day
(
M)
30
Days,
(
F)
35
days
periconce
ption
or
35
days
gestation
day
5
to
PND
1
(
M)
Clinical
pathology,
organ
weight,
sperm
analysis,
histopathology:
(
F)
maternal
weight,
reproductive
success,
pup
viability
and
growth
Paternal:
8.2
(
M);
10.8
(
F)

Reproductive/
de
velopmental:
8.2
(
M);
10.8
(
F)

(
Free­
standing
NOAEL)
ND
Smith
et
al.
(
1987)
Rat­

Long­
Evans
Hooded
Gavage
in
tricapyrlinb
50
mg/
kg/
day
Gestation
days
7
to
21
Maternal
weight,
reproductive
success,
pup
viability
and
growth
Maternal:
ND
Developmental:
ND
Maternal:
50
(
FEL
for
maternal
death;
decrease
maternal
weight
gain)

Development:
50
(
Decreased
litter
size,
decreased
fetal
weight)

a.
ND
=
not
determined.
b.
Due
to
the
demonstrated
ability
of
tricaprylin
to
enhance
developmental
toxicity
of
TCAN,
this
study
was
not
considered
in
derivation
of
the
Health
Advisories.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
24
Table
VIII­
3
Summary
of
Development
of
the
Health
Advisories
for
DBAN
Study
Critical
Effect
Critical
Effect
Level
Uncertainty
Factorsa
RfD
(
mg/
kg/
day)
Health
Advisory
(
mg/
L)

Ten­
day
NTP
(
2002)
Decreased
body
weight,
decreased
testes
weight,
and
testes
atrophy
12
mg/
kg/
day
(
NOAEL)
100
(
10H,
10A)
­
1
Longer­
term
NTP
(
2002)
Decreased
body
weight
11.3
mg/
kg/
day
(
NOAEL)
300
(
10H,
10A,
3D)
­
Child
0.4
Adult
1
Lifetime
NTP
(
2002)
Decreased
body
weight
11.3
mg/
kg/
day
(
NOAEL)
1000
(
10H,
10A,
3S,
3D)
0.01
0.08
a.
Areas
of
uncertainty
addressed
by
uncertainty
factors
are:
animal
to
human
extrapolation
(
A);
intrahuman
variability
and
protection
of
sensitive
subpopulations
(
H);
extrapolation
from
a
LOAEL
to
a
NOAEL(
L);
extrapolation
from
a
subchronic
to
chronic
exposure
(
S);
and
lack
of
a
complete
database
(
D)

B.
3
DCAN
B.
3.1
One­
Day
Health
Advisory
for
DCAN
The
oral
toxicity
data
for
DCAN
are
summarized
in
Table
VIII­
4.
Hayes
et
al.
(
1986)

identify
acute
oral
LD
50
values
for
DCAN
of
270
to
339
mg/
kg.
Clinical
signs
observed
include
ataxia,
depressed
respiration,
depressed
activity,
and
coma.
However,
LD
50
studies
are
not
suitable
for
the
development
of
one­
day
health
advisories
and
no
other
adequate
acute
studies
were
located
for
DCAN.
In
the
absence
of
such
data,
the
Ten­
day
HA
value
is
recommended
as
a
conservative
estimate
of
an
appropriate
One­
day
HA
value.

B.
3.2
Ten­
Day
Health
Advisory
for
DCAN
One
general
toxicity
study
of
suitable
duration
for
the
Ten­
day
HA
was
located.
Hayes
et
al.
(
1986)
conducted
a
14­
day
study
of
DCAN
toxicity
in
rats.
Body
weight
decreases
were
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
25
observed
in
both
males
and
females.
Males
were
more
sensitive
to
this
effect
than
females.
In
males,
a
decrease
in
body
weight
of
greater
than
10%
was
observed
at
45
and
90
mg/
kg/
day,

although
these
results
were
not
statistically
significant.
Several
serum
markers
for
organ
toxicity
were
increased
in
treated
animals.
Significantly
increased
SGPT
levels
in
females
at
90
mg/
kg/
day,
and
ALP
levels
at
90
mg/
kg/
day
in
males
and
at
45
and
90
mg/
kg/
day
in
females
were
reported,
possibly
indicative
of
hepatotoxicity.
Although
the
authors
did
not
consider
these
changes
to
be
compound­
related
adverse
effects
(
no
reason
provided),
these
changes
were
considered
adverse
for
this
assessment,
based
on
the
magnitude
of
the
changes,
and
the
supporting
data
for
DCAN
in
females
in
the
subchronic
study.
No
remarkable
findings
were
observed
at
necropsy;
however,
relative
liver
weight
was
significantly
increased
(
p

0.05)
in
male
and
female
rats.
The
observed
increase
in
serum
levels
of
hepatic
enzyme
activity
at
higher
doses
than
those
associated
with
liver
weight
gives
greater
weight
to
the
potential
toxicological
significance
of
the
liver
weight
changes,
even
though
the
absence
of
histopathology
data
makes
it
difficult
to
determine
conclusively
if
the
effects
were
adverse
at
low
doses.
Based
on
this
uncertainty,
both
decreased
body
weight
and
increased
relative
liver
weight
are
considered
toxicologically­
relevant
responses.
The
more
sensitive
of
these
endpoints
was
selected
as
the
critical
effect
for
this
study.
Therefore,
the
lowest
dose
tested
of
12
mg/
kg/
day
is
the
study
LOAEL
for
increased
relative
liver
weight
in
males,
and
no
NOAEL
is
determined.
BMD
modeling
was
conducted
for
decreased
body
weight
and
increased
relative
liver
weight
in
both
sexes
to
identify
alternative
critical
effect
levels
for
this
study.
A
BMDL
of
5
mg/
kg/
day
for
increased
relative
liver
weight
in
males
was
selected
as
the
most
appropriate
modeling
result
to
serve
as
the
basis
for
the
quantitative
dose­
response
assessment
(
see
Appendix
A).
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
26
As
a
follow­
up
to
their
earlier
single­
dose
study,
Smith
et
al.
(
1989)
reported
that
doses
of
25
to
45
mg/
kg/
day
administered
for
12
days
during
gestation
resulted
in
fetotoxicity
and
teratogenicity
in
rats.
However,
use
of
this
study
for
dose­
response
assessment
is
not
appropriate,
because
it
may
not
accurately
reflect
the
toxicity
of
DCAN
in
drinking
water.
This
conclusion
is
based
on
the
observation
of
embryotoxicity
of
the
tricaprylin
vehicle
in
this
study
and
later
work
by
this
laboratory
which
suggests
that
tricaprylin
may
act
synergistically
with
TCAN
to
enhance
developmental
toxicity
(
Christ
et
al.,
1996).

Based
on
increased
relative
liver
weight
in
male
rats
(
Hayes
et
al.,
1986)
the
Ten­
day
HA
was
calculated
as
shown
below
and
summarized
in
Table
VIII­
5.
An
uncertainty
factor
of
10
is
used
to
account
for
extrapolation
from
an
animal
study
and
an
uncertainty
factor
of
10
is
used
to
account
for
inter­
individual
variability
in
human
sensitivity.
An
additional
factor
of
3
was
used
to
account
for
extrapolation
from
a
minimal
LOAEL.
A
factor
of
10
was
not
used
since
the
adverse
effect
(
increased
relative
liver
weight)
was
of
marginal
severity
(
i.
e.
no
clinical
chemistry
findings
were
observed
at
this
dose).
This
additional
factor
was
not
used
for
derivation
of
the
Ten­
day
HA
when
the
BMDL
was
used
as
the
point
of
departure,
since
the
BMDL
often
approximates
a
NOAEL
as
indicated
by
the
lower
value
of
the
BMDL
for
the
same
effect
the
critical
study.
The
composite
uncertainty
factor
used
is
300
when
the
LOAEL
was
used
as
the
point
of
departure,

and
100
when
the
BMDL
was
used
as
the
point
of
departure.

Derivation
of
the
Ten­
day
HA
based
on
the
study
LOAEL.

(
12
mg/
kg/
day)
(
10
kg)
Ten­
day
HA
=
=
0.4
mg/
L
(
300)(
1
L/
day)
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
27
where:

12
mg/
kg/
day
=
LOAEL,
based
on
increased
relative
liver
weight
in
males
supported
by
clinical
chemistry
findings
at
higher
doses
in
rats
exposed
to
DCAN
by
gavage
for
14
days
(
Hayes
et
al.,
1986).

10
kg
=
assumed
body
weight
of
a
child.

300
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
minimal
LOAEL
in
animals,
and
inter­
individual
variability
in
humans.

1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

Derivation
of
the
Ten­
day
HA
based
on
the
study
BMDL.

(
5
mg/
kg/
day)
(
10
kg)
Ten­
day
HA
=
=
0.5
mg/
L
(
100)(
1
L/
day)

where:

5
mg/
kg/
day
=
BMDL,
based
on
increased
relative
liver
weight
in
males
supported
by
clinical
chemistry
findings
at
higher
doses
in
rats
exposed
to
DCAN
by
gavage
for
14
days
(
Hayes
et
al.,
1986).

10
kg
=
assumed
body
weight
of
a
child.

100
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
BMDL
in
animals,
and
inter­
individual
variability
in
humans.

1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

B.
3.3
Longer­
Term
Health
Advisory
for
DCAN
Only
one
study
of
suitable
duration
for
the
derivation
of
a
longer­
term
HA
was
located.

Hayes
et
al.
(
1986)
conducted
a
90­
day
subchronic
toxicity
study
in
rats.
Body
weight
was
significantly
decreased
in
high­
dose
male
and
female
rats
and
in
male
rats
at
33
mg/
kg/
day.

Relative
liver
weights
were
statistically
significantly
elevated
in
males
beginning
at
33
mg/
kg/
day
and
in
females
beginning
at
8
mg/
kg/
day.
However,
relative
liver
weight
increases
were
greater
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
28
than
10%
at
8
mg/
kg/
day
in
both
males
and
females,
and
therefore
the
increase
in
relative
liver
weight
at
this
dose
was
considered
toxicologically
relevant
for
both
sexes.
The
observed
increase
in
serum
levels
of
ALP
activity
in
the
subchronic
study,
and
the
increase
in
both
ALP
and
SGPT
observed
in
the
14­
day
study
support
the
toxicological
relevance
of
the
liver
weight
findings.
The
absence
of
histopathology
data
makes
it
difficult
to
determine
conclusively
if
the
effects
were
adverse
at
low
doses.
Based
on
this
uncertainty,
both
decreased
body
weight
and
increased
relative
liver
weight
are
considered
toxicologically­
relevant
responses.
The
more
sensitive
of
these
endpoints
was
selected
as
the
critical
effect.
Therefore,
the
lowest
dose
tested
of
8
mg/
kg/
day
is
the
study
LOAEL
for
increased
relative
liver
weight
in
males
and
females,
and
no
NOAEL
is
determined.
BMD
modeling
was
conducted
for
decreased
body
weight
and
increased
relative
liver
weight
in
both
sexes
to
identify
alternative
critical
effect
levels
for
this
study.
A
BMDL
of
4
mg/
kg/
day
for
increased
relative
liver
weight
in
males
was
selected
as
the
most
appropriate
modeling
result
to
serve
as
the
basis
for
the
quantitative
dose­
response
assessment,

using
a
benchmark
response
(
BMR)
of
a
one
standard
deviation
decrease
in
relative
liver
weight
(
see
Appendix
A).

Based
on
the
increased
relative
liver
weight
in
rats
(
Hayes
et
al.,
1986),
the
Longer­
term
HA
value
for
a
10­
kg
child
may
be
calculated
as
shown
below
and
summarized
in
Table
VIII­
5.

Derivation
of
the
health
advisories
are
shown
when
either
the
study
LOAEL
(
8
mg/
kg/
day)
or
BMDL
(
4
mg/
kg/
day)
is
selected
as
the
point
of
departure.
An
uncertainty
factor
of
10
is
used
to
account
for
extrapolation
from
an
animal
study
and
an
uncertainty
factor
of
10
is
used
to
account
for
human
variability
in
sensitivity,
in
the
absence
of
sufficient
data
to
depart
from
these
defaults.

An
uncertainty
factor
of
10
is
used
to
account
for
insufficiencies
in
the
database.
This
factor
was
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
29
chosen
because
only
one
subchronic
toxicity
study
in
a
single
species
was
identified
for
derivation
of
the
Longer­
term
HA
(
Table
VIII­
4).
The
absence
of
a
systemic
toxicity
study
of
suitable
duration
in
a
second
species,
the
lack
of
histopathology
data
in
the
existing
90­
day
study,
and
failure
to
investigate
effects
associated
with
thiocyanate
(
an
identified
metabolite)
or
cyanide
(
the
likely
precursor
of
thiocyanate),
such
as
thyroid
or
central
nervous
system
effects,
further
weakens
the
database.
In
addition,
no
adequate
studies
on
reproductive
or
developmental
toxicity
were
reported.
The
only
available
developmental
toxicity
study
testing
multiple
dose
levels
(
Smith
et
al.,
1989)
was
compromised
by
the
use
of
tricaprylin
as
the
solvent
vehicle
and
was
judged
as
inadequate
for
use
in
the
quantitative
dose­
response
assessment.
If
the
LOAEL
is
selected
as
the
point
of
departure,
an
additional
factor
of
3
is
used
to
account
from
extrapolation
from
a
LOAEL
for
minimally
adverse
liver
effects,
since
no
clinical
chemistry
changes
were
observed
at
this
dose
to
accompany
the
observed
increases
in
liver
weight.
The
composite
uncertainty
factor
used
is
3000
with
the
LOAEL
as
the
point
of
departure
and
1000
with
the
BMDL
as
the
point
of
departure,
based
on
full
factors
of
10
each
for
interspecies
extrapolation
and
inter­
individual
human
variability,
a
factor
of
10
for
database
insufficiencies,
and,
for
the
LOAEL
only,
a
partial
factor
of
3
for
a
minimal
LOAEL.

Derivation
of
the
Longer­
term
HA
based
on
the
study
LOAEL.

(
8
mg/
kg/
day)
(
10
kg)
Longer­
Term
HA
=
=
0.027
mg/
L
(
rounded
to
0.03
mg/
L)
(
3000)(
1
L/
day)

where:

8
mg/
kg/
day
=
LOAEL,
based
on
increased
relative
liver
weight
in
male
and
female
rats
exposed
to
DCAN
by
gavage
for
90
days
(
Hayes
et
al.,
1986).
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
30
10
kg
=
assumed
body
weight
of
a
child.

3000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
LOAEL
in
animals,
inter­
individual
variability
in
humans,
and
insufficiencies
in
the
database.

1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

The
Longer­
term
HA
for
a
70­
kg
adult
consuming
2
L/
day
of
water
is
calculated
as
follows:

(
8
mg/
kg/
day)
(
70
kg)
Longer­
Term
HA
=
=
0.093
mg/
L
(
rounded
to
0.09
mg/
L)
(
3000)(
2
L/
day)

where:

8
mg/
kg/
day
=
LOAEL,
based
on
increased
relative
liver
weight
in
male
and
female
rats
exposed
to
DCAN
by
gavage
for
90
days
(
Hayes
et
al.,
1986).

70
kg
=
assumed
body
weight
of
an
adult
3000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
LOAEL
in
animals,
inter­
individual
variability
in
humans,
and
insufficiencies
in
the
database.

2
L/
day
=
assumed
daily
water
consumption
by
a
70­
kg
adult
Derivation
of
the
Longer­
term
HA
based
on
the
study
BMDL.

(
4
mg/
kg/
day)
(
10
kg)
Longer­
Term
HA
=
=
0.04
mg/
L
(
1000)(
1
L/
day)

where:

4
mg/
kg/
day
=
BMDL,
based
on
increased
relative
liver
weight
in
male
rats
exposed
to
DCAN
by
gavage
for
90
days
(
Hayes
et
al.,
1986).

10
kg
=
assumed
body
weight
of
a
child.

1000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
BMDL
in
animals,
inter­
individual
variability
in
humans,
and
insufficiencies
in
the
database.

1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
31
The
Longer­
term
HA
for
a
70­
kg
adult
consuming
2
L/
day
of
water
is
calculated
as
follows:

(
4
mg/
kg/
day)
(
70
kg)
Longer­
Term
HA
=
=
0.14
mg/
L
(
rounded
to
0.1
mg/
L)
(
1000)(
2
L/
day)

where:

4
mg/
kg/
day
=
BMDL,
based
on
increased
relative
liver
weight
in
male
rats
exposed
to
DCAN
by
gavage
for
90
days
(
Hayes
et
al.,
1986).

70
kg
=
assumed
body
weight
of
an
adult
1000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
BMDL
in
animals,
inter­
individual
variability
in
humans,
and
insufficiencies
in
the
database.

2
L/
day
=
assumed
daily
water
consumption
by
a
70­
kg
adult
B.
3.4
Lifetime
Health
Advisory
for
DCAN
No
chronic
studies
of
DCAN
toxicity
were
located.
In
the
absence
of
such
data,
the
subchronic
(
90­
day)
LOAEL
of
8
mg/
kg/
day
or
the
BMDL
of
4
mg/
kg/
day
for
increased
relative
liver
weight
in
rats
reported
in
the
study
by
Hayes
et
al.
(
1986)
may
be
employed
to
derive
the
RfD
as
shown
below
and
summarized
in
Table
VIII­
5.
An
uncertainty
factor
of
10
is
used
to
account
for
extrapolation
from
an
animal
study
and
an
uncertainty
factor
of
10
is
used
to
account
for
inter­
individual
variability
in
human
sensitivity,
in
the
absence
of
sufficient
data
to
depart
from
these
defaults.

An
additional
factor
of
3
is
used
to
account
for
less­
than­
lifetime
exposure
for
DCAN.

This
factor
of
3
was
used
to
replace
the
default
factor
of
10
for
extrapolation
from
a
subchronic
study.
In
selecting
a
factor
of
3,
comparison
of
the
critical
effect
levels
between
the
14­
day
and
90­
day
studies
(
Hayes
et
al.,
1986)
is
not
helpful
since
the
low
dose
was
the
LOAEL
for
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
32
increased
liver
weight
in
both
cases.
An
added
complication
is
that
the
same
dose
levels
were
not
used
in
the
14­
day
and
90­
day
studies,
which
makes
comparison
of
the
severity
of
effects
with
duration
more
difficult.
However,
examination
of
the
liver
weight
changes
across
durations
suggests
that
the
uncertainty
factor
for
extrapolation
from
a
subchronic­
to­
chronic
study
would
be
adequately
accounted
by
an
uncertainty
factor
of
3.
This
conclusion
is
supported
by
the
observation
that
for
males
a
similar
degree
of
change
in
relative
liver
weight
(
increase
of
13%)

was
observed
over
a
1.5­
fold
change
in
dose
(
8
mg/
kg/
day
for
90
days
versus
12
mg/
kg/
day
for
14
days).
Furthermore,
in
females
the
increase
in
relative
liver
weight
was
roughly
two­
fold
greater
at
8
mg/
kg/
day
in
the
90­
day
study
than
at
12
mg/
kg­
day
in
the
14­
day
study,
suggesting
a
differences
in
responsiveness
of
3­
fold
on
an
equal
dose
basis
(
1.5­
fold
decrease
in
dose
x
2­
fold
increase
in
effect).
A
similar
type
of
evaluation
for
higher
doses
suggests
that
the
difference
in
the
magnitude
of
liver
weight
increases
is
not
likely
to
increase
by
ten­
fold,
with
increasing
exposure
duration,
although
some
increase
in
magnitude
of
the
effect
is
observed
with
longer­
term
exposure.
Finally,
the
BMDL
calculated
for
the
14­
day
study
(
5
mg/
kg/
day)
is
nearly
identical
to
the
BMDL
calculated
for
the
90­
day
study
(
4
mg/
kg/
day)
for
the
same
endpoint.
Since
the
BMDL
is
based
on
a
defined
change
in
mean
and
not
dependent
on
the
dose
levels,
the
similarity
of
the
BMDLs
across
study
duration
suggest
minimal
progression.

An
uncertainty
factor
of
10
is
used
to
account
for
insufficiencies
in
the
database.
This
factor
was
chosen
because
only
one
subchronic
toxicity
study
in
a
single
species
was
identified
for
derivation
of
the
Lifetime
HA
(
Table
VIII­
4).
The
absence
of
a
systemic
toxicity
study
of
suitable
duration
in
a
second
species,
the
lack
of
histopathology
data
in
the
existing
90­
day
study,
and
failure
to
investigate
effects
associated
with
thiocyanate
(
an
identified
metabolite)
or
cyanide
(
the
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
33
likely
precursor
of
thiocyanate),
such
as
thyroid
or
central
nervous
system
effects,
further
weakens
the
database.
In
addition,
no
sufficient
studies
on
reproductive
or
developmental
toxicity
were
reported.
The
only
available
multiple­
dose
developmental
toxicity
study
(
Smith
et.
al.,

1989)
was
compromised
by
the
use
of
tricaprylin
as
the
solvent
vehicle
and
was
judged
as
inadequate
for
use
in
the
quantitative
dose­
response
assessment.

If
the
LOAEL
is
selected
as
the
point
of
departure
an
additional
factor
of
3
is
used
to
account
from
extrapolation
from
a
LOAEL
for
minimally
adverse
liver
effects,
since
no
clinical
chemistry
changes
were
observed
at
this
dose
to
accompany
the
observed
increases
in
relative
liver
weight.
When
the
LOAEL
is
used
as
the
point
of
departure,
the
composite
uncertainty
factor
is
based
on
three
full
factors
of
10
and
two
partial
factors
of
3.
This
is
equivalent
to
four
full
areas
of
uncertainty.
Based
on
EPA
policy
(
Dourson,
1994)
the
composite
uncertainty
factor
for
four
individual
factors
of
10
is
3000,
reflecting
overlap
in
the
individual
factors.
Therefore,

the
composite
uncertainty
factor
is
3000
when
the
LOAEL
is
used
as
the
point
of
departure.

When
the
BMDL
is
used
as
the
point
of
departure,
the
composite
uncertainty
factor
is
based
on
three
full
factors
of
10
and
one
partial
factor
of
3.
Therefore,
the
composite
uncertainty
factor
is
also
3000
when
the
BMDL
is
used
as
the
point
of
departure.

Derivation
of
the
Lifetime
HA
based
on
the
study
LOAEL.

Step
1:
Determination
of
RfD
for
DCAN
(
8
mg/
kg/
day)
RfD
=
=
0.0027
mg/
kg/
day
(
rounded
to
0.003
mg/
kg/
day)
(
3000)

where:
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
34
8
mg/
kg/
day
=
LOAEL,
based
on
increased
relative
liver
weight
in
male
and
female
rats
exposed
to
DCAN
by
gavage
for
90
days
(
Hayes
et
al.,
1986).

3000
=
composite
uncertainty
factor
chosen
to
account
for
extrapolation
from
a
LOAEL
in
animals,
inter­
individual
variability
in
humans,
less­
than­
lifetime
exposure,
and
insufficiencies
in
the
database.

Step
2:
Determination
of
the
Drinking
Water
Equivalent
Level
(
DWEL)
for
DCAN
(
0.0027
mg/
kg/
day)
(
70
kg)
DWEL
=
=
0.094
mg/
L
(
rounded
to
0.09
mg/
L)
(
2
L/
day)

where:

0.0027
mg/
kg/
day
=
RfD.

70
kg
=
assumed
body
weight
of
an
adult.

2
L/
day
=
assumed
water
consumption
of
a
70­
kg
adult.

Step
3:
Determination
of
Lifetime
HA
for
DCAN
Lifetime
HA
=
(
0.094
mg/
L)
(
20%)
=
0.019
mg/
L
(
rounded
to
0.02
mg/
L)

where:

0.094
mg/
L
=
DWEL
20%
=
assumed
relative
source
contribution
from
water
Derivation
of
the
Lifetime
HA
based
on
the
study
BMDL.

Step
1:
Determination
of
RfD
for
DCAN
(
4
mg/
kg/
day)
RfD
=
=
0.0013
mg/
kg/
day
(
rounded
to
0.001
mg/
kg/
day)
(
3000)

where:

4
mg/
kg/
day
=
BMDL,
based
on
increased
relative
liver
weight
in
male
rats
exposed
to
DCAN
by
gavage
for
90
days
(
Hayes
et
al.,
1986).
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
35
3000
=
composite
uncertainty
factor
chosen
to
account
for
extrapolation
from
a
BMDL
in
animals,
inter­
individual
variability
in
humans,
less­
than­
lifetime
exposure,
and
insufficiencies
in
the
database.

Step
2:
Determination
of
the
Drinking
Water
Equivalent
Level
(
DWEL)
for
DCAN
(
0.0013
mg/
kg/
day)
(
70
kg)
DWEL
=
=
0.046
mg/
L
(
rounded
to
0.05
mg/
L)
(
2
L/
day)

where:

0.0013
mg/
kg/
day
=
RfD
(
before
rounding)

70
kg
=
assumed
body
weight
of
an
adult.

2
L/
day
=
assumed
water
consumption
of
a
70­
kg
adult.

Step
3:
Determination
of
Lifetime
HA
for
DCAN
Lifetime
HA
=
(
0.046
mg/
L)
(
20%)
=
0.0092
mg/
L
(
rounded
to
0.009
mg/
L)

where:

0.046
mg/
L
=
DWEL
20%
=
assumed
relative
source
contribution
from
water
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
36
Table
VIII­
4
Summary
of
Oral
Studies
of
DCAN
Toxicity
Reference
Species/
Strain
Route/
Dose
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

Hayes
et
al.
(
1986)
Mouse­

B6C3F1
Gavage
in
corn
oil
25
­
3,200
mg/
kg/
day
Acute
Lethality
NDa
LD50
=
270
(
M)
279
(
F)

Rat­

CD
Gavage
in
corn
oil
25
­
1,600
mg/
kg/
day
Acute
Lethality
ND
LD50
=
339
(
M)
330
(
F)

Rat­

CD
Gavage
in
corn
oil
0,
12,
23,
45,
90
mg/
kg/
day
14
Days
Body
weight,
organ
weight,
serum
chemistry,
hematology,
urinalysis,
gross
necropsy
ND
12
(
Increased
liver
weight)

Rat­

CD
Gavage
in
corn
oil
0,
8,
33,
65
mg/
kg/
day
90
Days
Body
weight,
organ
weight,
serum
chemistry,
hematology,
urinalysis,
gross
necropsy
ND
8
(
Increased
liver
weight)

Meier
et
al.
(
1985)
Mouse­

B6C3F1
Gavage
in
water
0,
12.5,
25,
or
50
mg/
kg/
day
5
Days
Sperm
head
abnormalities
50
mg/
kg
(
Free­
standing
NOAEL)
ND
Smith
et
al.
(
1987)
Rat­

Long­
Evans
Hooded
Gavage
in
tricapyrlinb
55
mg/
kg/
day
Gestation
days
7
to
21
Maternal
weight,
reproductive
success,
pup
viability
and
growth
ND
Maternal:
55
(
Decreased
maternal
weight)

Development:
55
(
Decreased
pregnancy
rate;
decreased
viable
litters;
increased
litters
resorbed;
decreased
fetal
weight)
Drinking
Water
Criteria
Document
for
Haloacetonitriles
Reference
Species/
Strain
Route/
Dose
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

EPA/
OW/
OST/
HECD
Final
Draft
VIII­
37
Smith
et
al.
(
1989)
Rat­

Long­
Evans
Hooded
Gavage
in
tricapyrlinb
0,
5,
15,
25,
45
mg/
kg/
day
Gestation
days
6
to
18
Maternal
weight,
reproductive
success,
pup
viability
and
growth
Maternal:
15
Developmental:
15
Maternal:
25
(
increased
liver
weight)

Development:
25
(
Increased
postimplantation
loss,
increased
soft­
tissue
malformations)

a.
ND
=
not
determined.
b.
Due
to
the
demonstrated
ability
of
tricaprylin
to
enhance
developmental
toxicity
of
TCAN,
this
study
was
considered
in
derivation
of
the
Health
Advisories.

Table
VIII­
5
Summary
of
Development
of
the
Health
Advisories
for
DCAN
Study
Critical
Effect
Critical
Effect
Level
Uncertainty
Factorsa
RfD
(
mg/
kg/
day)
Health
Advisory
(
mg/
L)

Ten­
day
Hayes
et
al.
(
1986)
Increased
relative
liver
weight
12
mg/
kg/
day
(
LOAEL)
300
(
10H,
10A,
3L)
­
0.4
Hayes
et
al.
(
1986)
Increased
relative
liver
weight
5
mg/
kg/
day
(
BMDL)
100
(
10H,
10A)
­
0.5
Longer­
term
Hayes
et
al.
(
1986)
Increased
relative
liver
weight
8
mg/
kg/
day
(
LOAEL)
3000
(
10H,
10A,
3L,
10D)
­
Child
0.03
Adult
0.09
Hayes
et
al.
(
1986)
Increased
relative
liver
weight
4
mg/
kg/
day
(
BMDL)
1000
(
10H,
10A,
10D)
­
Child
0.04
Adult
0.1
Lifetime
Hayes
et
al.
(
1986)
Increased
relative
liver
weight
8
mg/
kg/
day
(
LOAEL)
3000
(
10H,
10A,
3S,
3L,
10D)
0.003
0.02
Hayes
et
al.
(
1986)
Increased
relative
liver
weight
4
mg/
kg/
day
(
BMDL)
3000
(
10H,
10A,
3S,
10D)
0.001
0.009
a.
Areas
of
uncertainty
addressed
by
uncertainty
factors
are:
animal
to
human
extrapolation
(
A);
intrahuman
variability
and
protection
of
sensitive
subpopulations
(
H);
extrapolation
from
a
LOAEL
to
a
NOAEL(
L);
extrapolation
from
a
subchronic
to
chronic
exposure
(
S);
and
lack
of
a
complete
database
(
D)
Drinking
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38
B.
4
TCAN
B.
4.1
One­
Day
Health
Advisory
for
TCAN
The
oral
toxicity
data
for
TCAN
are
summarized
in
Table
VIII­
6.
An
oral
LD
50
of
360
mg/
kg
has
been
reported
(
Smyth
et
al.,
1962).
However,
LD
50
studies
are
not
suitable
for
the
development
of
one­
day
health
advisories
and
no
other
adequate
acute
studies
were
located
for
TCAN.
In
the
absence
of
suitable
acute
data,
the
Ten­
day
HA
value
is
recommended
as
a
conservative
estimate
of
the
One­
day
HA
value.

B.
4.2
Ten­
Day
Health
Advisory
for
TCAN
Two
developmental
studies
have
been
conducted
which
evaluate
the
short­
term
effects
of
TCAN.
In
Smith
et
al.
(
1988),
oral
exposure
of
pregnant
rats
to
TCAN
in
tricaprylin
on
gestation
days
6
through
18
resulted
in
significant
fetotoxic
and
teratogenic
effects
at
doses
of
7.5
mg/
kg/
day
or
higher.
Although
the
increased
incidence
of
teratogenic
effects
was
not
statistically
significant
at
a
dose
of
1
mg/
kg/
day,
the
authors
expressed
concern
that
this
might
be
of
biological
significance.
However,
further
evaluation
of
these
data
based
on
litter
incidences
did
not
reveal
a
treatment­
related
effect
at
1
mg/
kg/
day.
In
a
later
study
by
Christ
et
al.
(
1996),
oral
exposure
of
pregnant
rats
to
TCAN
in
corn
oil
on
gestation
days
6
to18
resulted
in
significantly
reduced
maternal
weight
gain
at
doses
of
35
mg/
kg/
day
and
higher.
Fetotoxic
and
teratogenic
effects
were
not
observed
until
doses
of
55
mg/
kg/
day
and
higher.
Because
of
the
confounding
effect
of
tricaprylin
toxicity,
as
identified
by
Christ
et
al.
(
1996),
exposure
to
TCAN
in
corn
oil
is
more
appropriate
basis
for
a
health
advisory.
Therefore,
the
Christ
et
al.
(
1996)
study
is
selected
as
the
critical
study
for
derivation
of
the
Ten­
day
HA
as
shown
below
and
summarized
in
Table
VIII­
7.

In
general,
developmental
toxicity
endpoints
arising
from
in
utero
exposure
have
limited
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
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OST/
HECD
Final
Draft
VIII­
39
application
as
the
basis
for
the
derivation
of
the
Ten­
day
HA,
since
the
Ten­
day
HA
is
based
on
water
consumption
by
children
directly,
and
not
from
maternal
exposure.
However,
when
the
critical
effect
is
a
general
systemic
endpoint
(
e.
g.,
decreased
maternal
weight
gain)
that
is
likely
to
be
relevant
to
exposed
children,
it
is
appropriate
to
use
these
data
to
derive
the
Ten­
day
HA.

Based
on
this
consideration,
the
study
by
Christ
et
al.
(
1996)
is
judged
as
an
adequate
basis
for
the
Ten­
day
HA.
Based
on
the
critical
effect
of
decreased
maternal
body
weight
gain,
the
NOAEL
is
15
mg/
kg/
day
and
the
LOAEL
is
35
mg/
kg/
day.
BMD
modeling
was
conducted
to
identify
alternative
critical
effect
levels
for
this
study.
A
BMDL
of
17
mg/
kg/
day
for
decreased
adjusted
maternal
weight
gain
was
selected
as
the
most
appropriate
modeling
result
to
serve
as
the
basis
for
the
quantitative
dose­
response
assessment
(
see
Appendix
A).

For
derivation
of
the
Ten­
day
HA,
an
uncertainty
factor
of
10
is
used
to
account
for
extrapolation
from
an
animal
study
and
an
uncertainty
factor
of
10
is
used
to
account
for
interindividual
variability
in
human
sensitivity,
in
the
absence
of
sufficient
data
to
depart
from
these
defaults.
The
composite
uncertainty
factor
used
is
100.

Derivation
of
the
Ten­
day
HA
based
on
the
study
NOAEL.

(
15
mg/
kg/
day)
(
10
kg)
Ten­
day
HA
=
=
1.5
mg/
L
(
rounded
to
2
mg/
L)
(
100)
(
1
L/
day)

where:

15
mg/
kg/
day
=
NOAEL,
based
on
the
absence
of
decreased
adjusted
maternal
weight
gain
in
pregnant
rats
exposed
by
gavage
on
days
6
to
18
of
gestation,
with
a
corresponding
LOAEL
of
35
mg/
kg/
day
(
Christ
et
al.,
1996).

10
kg
=
assumed
body
weight
of
a
child.
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Water
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Document
for
Haloacetonitriles
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HECD
Final
Draft
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40
100
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
NOAEL
from
a
study
in
animals,
and
inter­
individual
variability
in
humans.

1
L/
day
=
assumed
water
intake
by
a
10­
kg
child.

Derivation
of
the
Ten­
day
HA
based
on
the
study
BMDL.

(
17
mg/
kg/
day)
(
10
kg)
Ten­
day
HA
=
=
1.7
mg/
L
(
rounded
to
2
mg/
L)
(
100)
(
1
L/
day)

where:

17
mg/
kg/
day
=
BMDL,
based
on
the
absence
of
decreased
adjusted
maternal
weight
gain
in
pregnant
rats
exposed
by
gavage
on
days
6
to
18
of
gestation
(
Christ
et
al.,
1996).

10
kg
=
assumed
body
weight
of
a
child.

100
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
NOAEL
from
a
study
in
animals,
and
inter­
individual
variability
in
humans.

1
L/
day
=
assumed
water
intake
by
a
10­
kg
child.

B.
4.3
Longer­
Term
and
Lifetime
Health
Advisories
for
TCAN
No
data
on
the
effects
of
longer­
term
or
chronic
exposure
to
TCAN
were
located.
In
the
absence
of
suitable
data,
no
values
can
be
derived
for
the
Longer­
term
or
Lifetime
HAs
for
TCAN.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
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HECD
Final
Draft
VIII­
41
Table
VIII­
6
Summary
of
Oral
Studies
of
TCAN
Toxicity
Reference
Species/
Strain
Route
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

Smyth
et
al.
(
1962)
Rat­

Wistar
Gavage
0.19
­
0.32
mg/
kg/
day
Acute
Lethality
NDa
LD50
=
360
Meier
et
al.
(
1985)
Mouse­

B6C3F1
Gavage
in
water
0,
12.5,
25,
or
50
mg/
kg/
day
5
Days
Sperm
head
abnormalities
50
mg/
kg
(
Freestanding
NOAEL)
ND
Smith
et
al.
(
1987)
Rat
Long­
Evans
Hooded
Gavage
in
tricaprylinb
55
mg/
kg/
day
Days
7
to
21
of
gestation
Maternal
weight,
reproductive
success,
pup
viability
and
growth
Maternal:
ND
Developmental:
ND
Maternal:
55
(
FEL
for
maternal
death;
decrease
maternal
weight
gain)

Development:
55
(
Decreased
pregnancy
rate;
decreased
viable
litters;
increased
litters
resorbed;
decreased
fetal
weight)

Smith
et
al.
(
1988)
Rat
Long­
Evans
Hooded
Gavage
in
tricaprylinb
0,
1,
7.5,
15,
35,
55
mg/
kg/
day
Days
6
to
18
of
gestation
Maternal
weight,
reproductive
success,
pup
viability
and
growth,
malformations
Maternal:
35
Developmental:
1
Maternal:
55
(
FEL
for
maternal
death;
decrease
maternal
weight
gain)

Developmental:
7.5
(
Increased
full­
liter
resorptions;
increased
cardiovascular
malformations)

Christ
et
al.
(
1996)
Rat
Long­
Evans
Gavage
in
corn
oilc
0,
15,
35,
55,
75
mg/
kg/
day
Days
6
to
18
of
gestation
Maternal
body
and
organ
weight,
reproductive
success,
pup
viability
and
growth,
malformations
Maternal:
15
Developmental:
35
Maternal:
35
(
Decreased
maternal
weight
gain;
organ
weight
changes)

Development:
55
(
increased
postimplantation
loss,
cardiovascular
and
cranio­
facial
malformations;
decreased
live
fetuses
per
litter,
fetal
body
weight,
crown­
rump
length.

a.
ND
=
not
determined
b.
Due
to
the
demonstrated
ability
of
tricaprylin
to
enhance
developmental
toxicity
of
TCAN,
this
study
was
not
considered
in
derivation
of
the
Health
Advisories.
c.
Only
data
relating
to
the
corn
oil
control
are
reported
in
the
table,
since
the
developmental
toxicity
reported
in
the
groups
administered
tricaprylin
were
not
considered
in
derivation
of
the
Health
Advisories.
Drinking
Water
Criteria
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for
Haloacetonitriles
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Table
VIII­
7.
Summary
of
Development
of
the
Health
Advisories
for
TCAN.

Study
Critical
Effect
Critical
Effect
Level
Uncertainty
Factorsa
RfD
(
mg/
kg/
day)
Health
Advisory
(
mg/
L)

Ten­
day
Christ
et
al.
(
1996)
Decreased
adjusted
maternal
body
weight
gain
15
mg/
kg/
day
(
NOAEL)
100
(
10H,
10A)
­
2
Christ
et
al.
(
1996)
Decreased
adjusted
maternal
body
weight
gain
17
mg/
kg/
day
(
BMDL)
100
(
10H,
10A)
­
2
Longer­
term
Inadequate
data
to
derive
Health
Advisory
Lifetime
Inadequate
data
to
derive
Health
Advisory
a.
Areas
of
uncertainty
addressed
by
uncertainty
factors
are:
animal
to
human
extrapolation
(
A);
intrahuman
variability
and
protection
of
sensitive
subpopulations
(
H).

C.
Carcinogenic
Effects
No
epidemiological
studies
have
evaluated
directly
the
carcinogenic
potential
of
HANs
in
humans.
Rather,
studies
have
evaluated
the
carcinogenic
potential
of
chlorinated
versus
unchlorinated
drinking
water
or
the
presence
of
trihalomethanes
as
a
marker
of
chlorination
byproducts
(
IARC,
1999;
Mills
et
al.,
1998).
No
standard
cancer
bioassays
of
HANs
have
been
done
in
animals,
although
DBAN
is
on
test
for
a
full
bioassay
as
part
of
the
National
Toxicology
Program
(
NTP,
2002).
Limited
short­
term
exposure
data
from
the
mouse
skin
assay
(
Bull
et
al.,

1985)
and
the
mouse
lung
assay
(
Bull
and
Robinson,
1985)
indicate
that
all
four
compounds
(
BCAN,
DBAN,
and
TCAN)
may
be
tumorigenic,
although
DCAN,
DBAN,
and
TCAN
were
reported
to
be
negative
in
the
rat
liver
GGT­
foci
assay
(
Herren­
Freund
and
Pereira,
1986).
QSTR
predictions
of
the
carcinogenicity
of
the
HANs
have
produced
mixed
results
(
Moudgal
et
al.,
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
VIII­
43
2000).
For
example,
BCAN
was
predicted
as
positive
in
male
and
female
mice,
but
negative
in
both
sexes
of
rats.
DBAN
was
negative
in
both
sexes
of
mice,
and
female
rats,
but
indeterminate
in
male
rats.
No
carcinogenicity
predictions
for
TCAN
or
DCAN
were
included
in
the
study.

Taken
together,
the
screening
bioassays
and
QSTR
predictions
provide
at
least
limited
indications
of
potential
carcinogenicity
of
HANs,
although
the
existing
data
are
not
adequate
to
demonstrate
positive
carcinogenicity
in
animals.

The
HANs
or
their
metabolites
are
reactive
compounds
that
can
bind
macromolecules
including
DNA
and
proteins
(
Daniel
et
al.,
1986;
Lin
et
al.,
1992).
Glutathione
conjugation
may
be
an
important
cellular
protection
against
these
reactive
compounds
(
Ahmed
et
al.,
1991),

although
no
glutathione
conjugates
or
their
metabolites
have
been
identified.
The
genotoxicity
of
each
of
the
HAN
compounds
BCAN,
DBAN,
DCAN,
and
TCAN
have
been
evaluated
in
at
least
one,
and
in
most
cases
a
variety
of
different
assays.
Although
some
of
the
data
have
provided
contradictory
results,
all
of
the
tested
compounds
appear
to
have
some
capacity
to
induce
genotoxic
effects.
For
example,
each
compound
has
been
found
to
generate
a
positive
result
in
at
least
one
reported
Salmonella/
microsome
assay,
except
DBAN.
In
addition,
while
not
uniformly
consistent,
a
variety
of
other
assays
for
DNA
damage
(
i.
e.
DNA
strand
breaks)
or
responses
to
DNA
damage
(
sister
chromatid
assays,
tests
for
gene
recombination
in
yeast,
and
SOS
chromotest)
have
yielded
positive
results
for
some
of
the
HANs.
Overall,
these
data
suggest
that
HANs
induce
genotoxicity
through
direct
interactions
with
DNA.
Evidence
for
direct
chromosome
effects
are
weaker,
with
inconsistent
results
reported
in
the
limited
number
of
studies
that
evaluated
formation
of
micronuclei
and
aneuploidy.
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The
International
Agency
for
Research
on
Cancer
(
IARC,
1999)
reviewed
the
data
for
HANs
and
found
that
there
is
inadequate
evidence
for
carcinogenicity
in
experimental
animals
for
all
compounds.
As
a
results,
IARC
classified
these
compounds
as
"
not
classifiable
as
to
its
carcinogenicity
to
humans"
(
Group
3).

Following
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1986),

BCAN,
DBAN,
DCAN,
and
TCAN
are
appropriately
classified
as
Group
D
­
Not
Classifiable
as
to
Human
Carcinogenicity.
This
classification
is
appropriate
when
there
is
inadequate
evidence
of
carcinogenicity
in
humans
or
animals.
Following
EPA's
Draft
1999
Guidelines
for
Carcinogen
Risk
Assessment
(
U.
S.
EPA,
1999),
the
data
for
the
HANs
can
best
be
described
as
Data
Are
Inadequate
for
an
Assessment
of
Human
Carcinogenic
Potential.

D.
Characterization
of
Uncertainties
and
Data
Gaps
The
available
data
are
very
limited
for
the
HANs
included
in
this
Criteria
Document.

Adequate
human
data
are
not
available
to
evaluate
noncancer
or
cancer
effects.
Full
chronic
toxicity
and
carcinogenicity
studies
in
animals
were
not
available
for
any
of
the
HANs.
The
absence
of
adequate
carcinogenicity
testing
is
a
significant
data
gap,
particularly
in
light
of
the
mixed
results
in
screening
bioassays,
genotoxicity
testing,
and
QSTR
modeling,
which
indicate
at
least
some
potential
for
HANs
to
induce
tumorigenic
responses.
Available
subchronic
studies
for
DCAN
did
not
include
a
full
histopathology
evaluation
of
relevant
tissues.
The
absence
of
histopathology
evaluations
following
longer­
term
exposures
resulted
in
significant
uncertainties
in
the
assessment
of
noncancer
toxicity,
since
the
adversity
of
the
increases
in
liver
weight
induced
Drinking
Water
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Final
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by
DCAN
(
Hayes
et
al.,
1986)
could
not
be
substantiated,
and
detailed
evaluations
of
known
targets
for
the
oxidative
pathway
metabolites
cyanide
and
thiocyanate
was
not
possible.

Although
the
developmental
toxicity
of
HANs
has
been
evaluated
in
numerous
animal
studies,
the
confounding
effects
of
the
tricaprylin
solvent
vehicle
that
was
used
in
the
earlier
studies
greatly
limits
the
usefulness
of
most
of
the
data.
The
mechanism
by
which
tricaprylin
impacted
the
developmental
toxicity
of
the
HANs
in
these
studies
remains
unclear.
Due
to
confounding
by
tricaprylin,
adequate
developmental
toxicity
data
are
available
only
for
TCAN
(
Christ
et
al.,
1996).
No
multi­
generation
reproductive
studies
were
available
for
any
of
the
HANs.
Only
one
study
was
available
that
evaluated
the
developmental
effects
of
HANs
in
humans
(
Klotz
and
Pyrch,
1999),
and
no
association
was
observed
between
HANs
exposure
and
developmental
effects.
The
potential
reproductive
and
developmental
toxicity
of
the
HANs
remains
a
significant
area
of
uncertainty
in
the
current
assessment,
and
represents
a
major
data
gap
in
light
of
the
reproductive
and
developmental
effects
attributed
to
other
disinfectant
byproducts
in
humans
(
Neiuwenhuijsen
et
al.,
2000).

The
available
data
on
HAN
toxicokinetics
and
toxicodymanics
was
not
sufficient
to
move
away
from
default
uncertainty
factor
values
for
extrapolation
from
animal
studies
or
for
interindividual
variability
in
human
sensitivity.
Basic
research
on
toxicokinetics
is
needed
for
most
of
the
HANs;
only
DBAN
(
NTP,
2002)
and
DCAN
have
been
studied
in
detail
(
Roby
et
al.,
1986),

and
only
the
study
for
DCAN
was
available
for
review
at
the
time
this
document
was
prepared.

In
particular,
further
understanding
of
HAN
metabolism
is
needed.
Research
in
this
area
could
clarify
the
relative
contribution
of
glutathione
conjugation
and
oxidative
metabolism
pathways
to
Drinking
Water
Criteria
Document
for
Haloacetonitriles
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Final
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the
observed
spectrum
of
toxicity
for
HANs.
Identification
of
the
toxic
moiety
and
the
GST
or
CYP
isoforms
important
in
catalyzing
HAN
metabolism
could
contribute
to
characterization
of
potential
human
susceptibility
based
on
age,
gender,
or
genetic
predisposition.

Based
on
the
clear
limitations
in
the
database
and
gaps
in
understanding
of
the
mechanisms
of
toxicity
for
HANs,
the
derived
RfD
and
HA
values
are
best
characterized
as
low
in
confidence.
