Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
1
Chapter
V.
Health
Effects
in
Animals
Limited
toxicity
testing
data
are
available
for
the
HANs.
Tables
V­
15
through
V­
18
at
the
end
of
the
chapter
provide
a
summary
of
the
toxicity
studies
for
BCAN,
DBAN,
DCAN,
and
TCAN
for
noncarcinogenic
endpoints.

A.
Short­
Term
Exposures
Acute
oral
LD
50
values
for
DBAN,
DCAN,
and
TCAN
in
rodents
have
been
reported
by
several
investigators,
as
summarized
in
Table
V­
1.
Acute
oral
LD
50
values
for
DBAN
and
DCAN
were
reported
to
range
from
245
to
361
mg/
kg
in
male
and
female
CD
rats
and
mice
given
single
oral
doses
dissolved
in
corn
oil.
Ataxia,
depressed
respiration,
depressed
activity,
and
coma
preceded
death.
No
consistent,
compound­
related,
gross
pathological
effects
were
observed
at
necropsy
(
Hayes
et
al.,
1986).
In
a
limited
report
on
the
acute
toxicity
of
DBAN,
groups
of
20
rats
and
mice
(
sex
and
strain
not
specified)
were
given
single
gavage
doses
of
10%
DBAN
in
corn
oil
ranging
from
25
to
1,600
mg/
kg
in
rats
and
from
25
to
3,200
mg/
kg
in
mice.
The
LD
50
was
estimated
as
50
to
100
mg/
kg
in
rats
and
50
mg/
kg
in
mice
with
slight
to
moderate
convulsions
the
only
symptom
reported
(
Eastman
Kodak
Co.,
1992).
Smyth
et
al.
(
1962)
reported
an
oral
LD
50
of
0.25
mL/
kg
(
360
mg/
kg)
in
male
Wistar
rats
for
TCAN
(
vehicle
was
not
specified).
Drinking
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­
2
Table
V­
1.
Acute
Oral
Lethality
of
Haloacetonitriles.

Reference
Species
LD50
(
mg/
kg)
Male
Female
DBAN
Hayes
et
al.
(
1986)
Mouse
289
(
253­
324)
a
303
(
269­
342)

Hayes
et
al.
(
1986)
Rat
245
(
210­
286)
361
(
320­
410)

Eastman
Kodak
Co.,
1992
Mouse
50
(
sex
not
specified)

Eastman
Kodak
Co.,
1992
Rat
50
­
100
(
sex
not
specified)

DCAN
Hayes
et
al.
(
1986)
Mouse
270
(
241
­
303)
279
(
263
­
296)

Hayes
et
al.
(
1986)
Rat
339
(
298
­
387)
330
(
300
­
500)

TCAN
Smyth
et
al.
(
1962)
Rat
360
a95%
Confidence
limits.

Lin
and
Guion
(
1989)
reported
clinical
signs
of
toxicity
in
a
mechanistic
study
to
evaluate
the
ability
of
HANs
to
deplete
liver
GSH
and
inhibit
GST
activity.
Male
Fischer
344
rats
were
administered
single
gavage
doses
of
0.75
mmol/
kg
BCAN
(
116
mg/
kg),
DBAN
(
149
mg/
kg),

DCAN
(
82
mg/
kg),
or
TCAN
(
108
mg/
kg)
in
tricaprylin.
These
doses
represented
10
to
30%
of
the
reported
LD
50
for
the
individual
compounds.
The
incidence
of
deaths
at
18
hours
was
2
of
20
for
TCAN
and
1
of
10
for
DBAN.
No
deaths
were
reported
for
the
other
HANs.
Overt
signs
of
acute
toxicity,
including
gasping
and
salivation,
were
observed
for
TCAN
and
DBAN.
Drinking
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­
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The
acute
toxicity
of
DBAN
has
also
been
evaluated
by
the
inhalation
route.
Sprague­

Dawley
rats
(
10
animals/
sex/
dose)
were
exposed
for
6
hours
to
air
concentrations
of
3.5,
9.5,

17.9,
or
41
ppm
DBAN
(
28.5,
77.3,
145.6,
or
333.5
mg/
m3)(
Dow
Chemical
Co.,
1992).
Slight
transient
eye
and
nasal
irritation
was
noted
at
the
lowest
concentration
and
increased
in
severity
with
increasing
concentration.
In
addition,
respiratory
tract
irritation
was
observed
at
9.5
ppm
and
increased
in
severity
with
increasing
concentration.
Three
males
and
two
females
did
not
survive
following
exposure
to
9.5
ppm;
all
males
and
6
females
in
the
17.9
ppm
exposure
group
died.
The
LC
50
was
reported
as
9.6
ppm
for
males
and
14.4
ppm
for
females,
respectively,

calculated
using
a
moving
average
method.
In
the
low
exposure
group,
no
treatment­
related
effects
were
observed
at
necropsy.
In
the
9.5
ppm
group,
nasal
irritation
and
distended
stomach
were
the
principal
findings
in
the
animals
that
did
not
survive;
no
effects
were
observed
in
the
surviving
animals.
In
the
two
high­
exposure
groups,
necropsy
revealed
distended
stomachs,
and
congested
liver
and
lungs.
No
treatment­
related
lesions
were
observed
following
gross
pathological
examination
in
the
four
females
that
survived
the
17.9
ppm
exposure.
In
the
study
by
Smyth
et
al.
(
1962),
no
mortality
was
observed
following
a
single
4­
hour
exposure
to
125
ppm
(
738
mg/
m3)
of
TCAN
in
a
group
of
six
albino
rats
(
sex
and
strain
not
specified),
indicating
that
the
4­
hour
LC
50
would
be
well
above
125
ppm.
Data
were
not
reported
for
end
points
other
than
mortality
in
this
study.
Although
the
different
exposure
durations
and
strains
used
preclude
a
direct
comparison,
the
finding
that
the
4­
hour
LC
50
for
TCAN
was
likely
much
higher
than
the
6­

hour
LC
50
for
DBAN
suggests
that
DBAN
has
higher
acute
inhalation
toxicity.
No
inhalation
studies
were
identified
for
BCAN
or
DCAN.
Drinking
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Document
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In
a
poorly­
reported
skin
irritation
study,
undiluted
DBAN
was
applied
to
the
skin
of
guinea
pigs
(
3
animals/
dose)
at
doses
ranging
from
0.01
to
1.0
mL/
kg
(
23
to
2,300
mg/
kg)(
Eastman
Kodak
Co.,
1992).
The
estimated
LD
50
was
0.1
to
1.0
mL/
kg
(
230
to
2,300
mg/
kg).
Irritant
symptoms
were
recorded
as
moderate
to
gross
edema,
with
necrosis
of
the
entire
patch
area
and
hemorrhagic
periphery
at
24
hours.
At
1
week,
the
authors
reported
heavy
eschar
that
was
broken
at
the
edges
with
secondary
eschar
beneath.
At
2
weeks,
small
secondary
eschar
was
noted,
with
heavy
scarring
and
alopecia.
The
LD
50
value
for
a
single
dermal
application
of
TCAN
to
male
New
Zealand
rabbits
was
reported
to
be
0.90
mL/
kg
(
1,296
mg/
kg)(
Smyth
et
al.,

1962).
TCAN
is
extremely
irritating;
0.5
mL
of
a
1%
solution
produced
a
severe
eye
burn
in
rabbits,
and
0.01
mL
of
the
material
caused
necrosis
when
applied
to
the
clipped
skin
of
rabbits.

The
study
descriptions
did
not
indicate
if
the
TCAN
was
undiluted
or
applied
in
a
solvent
vehicle.

No
dermal
studies
were
identified
for
BCAN
or
DCAN.

In
addition
to
acute
lethality
and
irritation,
other
effects
of
short­
term
exposure
to
DBAN
and
DCAN
have
been
evaluated.
The
evaluation
of
several
endpoints
for
DBAN
toxicity
have
been
conducted
or
are
being
planned.
DBAN
has
been
selected
for
a
neurotoxicity
study
(
NTP,

2002),
has
been
tested
in
a
short­
term
reproductive
and
developmental
toxicity
screening
study
(
R.
O.
W.
Sciences,
1997),
and
has
been
evaluated
in
14­
day
dose­
range
finding
studies
(
NTP,

2002).
These
latter
two
studies
are
described
in
this
section.
Drinking
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­
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The
short­
term
effects
of
DBAN
were
evaluated
as
part
of
dose­
range
finding
studies
conducted
for
a
reproductive
and
developmental
toxicity
screening
study
(
R.
O.
W.
Sciences,

1997).
In
the
initial
dose­
range
finding
study,
male
and
female
Sprague­
Dawley
rats
(
5/
sex/
dose)

were
exposed
to
drinking
water
containing
0,
250,
500,
1000,
or
2000
ppm
DBAN
(
doses
resulting
from
these
exposures
were
not
reported
by
the
study
authors).
Significant
decreases
in
body
weight
relative
to
controls
and
decreased
food
consumption
were
observed
within
4
days
of
exposure,
beginning
at
250
ppm
in
males
and
500
ppm
in
females.
Water
consumption
was
significantly
decreased
at
the
lowest
concentration
tested
(
250
ppm)
in
males
and
females.
The
onset
of
these
effects
led
the
authors
to
terminate
the
study,
since
the
lowest
concentration
used
in
this
first
range­
finding
study
was
judged
to
be
too
high
for
use
in
the
main
study.
Therefore,

the
rats
were
allowed
to
recover
to
control
body
weights,
and
were
exposed
to
a
lower
set
of
range­
finding
concentrations
of
0,
7,
20,
70,
or
200
ppm
in
drinking
water
for
2
weeks.
The
rats
were
evaluated
for
clinical
signs
of
toxicity,
body
weight,
and
feed
and
water
consumption.

Based
on
measured
water
consumption
and
body
weights,
the
estimated
dose
levels
in
the
second
range­
finding
study
were
0,
0.7,
2.2,
5.8,
and
13.2
mg/
kg/
day
for
the
males
and
0,
0.8,
2.4,
6.8,

and
17.9
mg/
kg/
day
for
the
females.
No
clinical
signs
of
toxicity
were
observed
in
any
of
the
exposure
groups.
A
significant
decrease
in
body
weights
of
roughly
10%
was
observed
in
males
exposed
to
20
or
200
ppm
DBAN
after
4
days
of
treatment.
However,
no
similar
decrease
was
observed
in
the
males
exposed
to
70
ppm,
and
the
mean
final
weight
and
total
weight
gain
at
the
end
of
the
study
were
not
affected
in
any
dose
group.
No
significant
effects
on
feed
consumption
were
observed.
Water
consumption
was
significantly
decreased
(
p

0.05)
on
day
11
in
females
Drinking
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­
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exposed
to
70
ppm
(
to
71%
of
controls).
Water
consumption
was
significantly
decreased
at
each
time
point
examined
(
study
days
4,
8,
11,
and
15)
for
males
and
females
in
the
200­
ppm
group.

The
maximum
decrease
in
water
consumption
was
to
50%
of
controls
for
males
and
to
67%
of
controls
for
females.
The
absence
of
clinical
signs
of
toxicity
or
body
weight
changes
indicates
that
the
highest
concentration
tested
(
200
ppm
in
the
second
range­
finding
study)
is
a
study
NOAEL.
Decreased
water
consumption
is
not
judged
to
be
an
adverse
toxicological
effect,
as
this
result
may
reflect
poor
palatability
of
the
DBAN­
containing
water.
Based
on
these
considerations,
the
study
NOAEL
is
the
highest
dose
tested
(
13.2
mg/
kg/
day
in
males;
17.9
mg/
kg/
day
in
females)
and
no
LOAEL
was
determined.

In
the
main
study
(
described
in
more
detail
in
the
Reproductive
and
Developmental
Toxicity
Section),
Sprague­
Dawley
rats
were
given
0,
15,
50,
or
150
ppm
DBAN
in
their
drinking
water.
Male
rats
(
10
animals/
dose)
were
given
treated
water
on
study
days
6
through
35,
and
were
then
examined
for
clinical
pathology
(
hematology
and
clinical
chemistry),
body
and
organ
weight
changes
(
liver,
right
kidney,
spleen,
thymus,
right
testis,
right
epididymis,
right
cauda
epididymis),
sperm
analyses,
and
histopathology.
Estimated
DBAN
doses
in
males
were
0,
1.4,

3.3,
and
8.2
mg/
kg/
day
as
calculated
by
the
study
authors
from
body
weight
and
water
consumption
data.
Separate
groups
of
female
rats
were
exposed
to
the
same
concentrations
as
males
on
study
days
1­
34
(
Group
A)
or
from
gestation
day
6
to
postnatal
day
1
(
Group
B).

Estimated
DBAN
doses
were
0,
1.8,
5.1,
and
10.9
mg/
kg/
day
for
Group
A
females,
and
0,
1.9,

5.3,
and
10.8
mg/
kg/
day
for
Group
B
females.
For
both
groups
of
females,
clinical
signs
of
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toxicity,
body
weight,
and
feed
and
water
consumption
were
determined
at
various
intervals,
and
general
pathology
was
evaluated
at
necropsy.
No
compound­
related
effect
was
observed
for
any
of
these
parameters.
The
only
biologically­
significant
effect
that
was
treatment
related
for
males
or
females
was
a
decrease
in
water
consumption
in
the
mid­
and
high­
concentration
groups,
which
might
have
been
related
to
decreased
water
palatability.
Since
the
males
were
examined
for
more
sensitive
endpoints
than
females
(
i.
e.
clinical
chemistry
and
histopathology
were
evaluated),
the
NOAEL
for
systemic
effects
in
this
study
was
8.2
mg/
kg/
day,
the
highest
dose
level
in
males,
and
no
LOAEL
was
identified.

The
subacute
toxicity
of
DBAN
has
also
been
evaluated
by
the
NTP
(
2002)
in
B6C3F1
mice
and
F344
rats
as
part
of
initial
dose­
range
finding
studies
in
support
of
chronic
exposure
studies
that
are
currently
in
progress.
For
the
mouse
study,
DBAN
was
administered
in
drinking
water
for
14
days
to
male
and
female
B6C3F1
mice
(
5/
sex/
dose)
at
concentrations
of
0,
12.5,
25,

50,
100,
or
200
mg/
L.
The
corresponding
doses
reported
by
the
study
authors
were
0,
2.1,
4.3,

8.2,
14.7,
and
21.4
mg/
kg/
day
for
males
and
2.0,
3.3,
10.0,
13.9,
and
21.6
mg/
kg/
day
for
females.

Animals
were
observed
for
clinical
signs
of
toxicity,
as
well
as
body
weight,
and
organ
weight
and
pathology.
In
addition,
liver
glutathione­
S­
transferase
(
GST)
activity
was
measured.
The
only
treatment­
related
effect
was
a
decrease
in
water
consumption
in
both
males
and
females.
The
decrease
in
water
consumption
was
concentration­
related,
decreasing
to
58%
of
controls
for
males
and
54%
controls
for
the
200
mg/
L
group
mice.
Since
the
only
effect
observed
was
a
concentration­
related
decrease
in
water
intake,
which
could
reflect
poor
water
palatability,
the
Drinking
Water
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Document
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­
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high
doses
of
21.4
mg/
kg/
day
for
males
and
21.6
mg/
kg/
day
for
females
is
considered
a
NOAEL.

No
LOAEL
is
identified.

For
the
rat
study,
DBAN
was
administered
in
drinking
water
for
14
days
to
male
and
female
F344
rats
(
5/
sex/
dose)
at
concentrations
of
0,
12.5,
25,
50,
100,
or
200
mg/
L.
The
corresponding
doses
reported
by
the
study
authors
were
0,
2,
3,
7,
12,
and
18
mg/
kg/
day
for
males
and
2,
4,
7,
12,
and
19
mg/
kg/
day
for
females.
Animals
were
observed
for
clinical
signs
of
toxicity,
as
well
as
body
weight,
and
organ
weight
and
pathology.
In
addition,
liver
glutathione­

S­
transferase
(
GST)
activity
was
measured.
A
concentration­
related
decrease
in
water
consumption
was
observed
for
both
males
and
females.
Water
consumption
decreased
to
60%
of
controls
for
males
and
61%
controls
for
females
in
the
200
mg/
L
group
mice.
No
effects
other
than
decreased
water
consumption
were
noted
in
females.
However,
in
males
DBAN
exposure
caused
a
decrease
in
body
weight
gain
and
terminal
body
weight
that
was
judged
to
be
toxicologically­
significant
only
at
the
high
dose.
The
reported
body
weight
gains
for
males
were
61.1,
66.3,
66.0,
65.2,
56.4,
and
34.0
grams
for
the
control
and
increasing
dose­
groups,

respectively.
Terminal
body
weights
as
percent
of
controls
were
100,
104.5,
104.3,
101.3,
98.7,

82.7%
for
the
control
and
increasing
dose­
groups,
respectively.
Significantly
decreased
testes
weights
were
observed
in
the
high
dose­
group
males.
This
finding
was
accompanied
by
testicular
atrophy
in
2
of
5
males
in
this
dose
group.
Elevated
liver
GST
activity
(
126%
of
controls)
was
also
reported
for
high
dose
males.
Based
on
decreased
body
weight
and
decreased
testes
weight
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
9
and
pathology
in
males,
the
NOAEL
for
this
study
is
12
mg/
kg/
day
and
the
LOAEL
is
18
mg/
kg/
day.

Hayes
et
al.
(
1986)
investigated
the
subacute
(
14­
day)
toxicity
of
DBAN
in
adult
CD
rats.

The
chemical
was
administered
daily
by
gavage
in
corn
oil
at
doses
of
0,
23,
45,
90,
or
180
mg/
kg/
day
(
10
animals/
sex/
dose).
Endpoints
assessed
included
mortality,
body
weight,
organ
weight,
serum
chemistry,
hematology,
urinalysis,
and
gross
necropsy,
as
shown
in
Tables
V­
2
and
V­
3.
Histopathological
studies
were
not
conducted.
Also,
food
and
water
consumption
rates
were
not
measured.
There
was
100%
mortality
at
180
mg/
kg/
day.
At
90
mg/
kg/
day,
40%
of
the
males
and
20%
of
the
females
were
dead
by
day
14.
No
consistent,
compound­
related
and
dosedependent
adverse
effects
were
apparent
in
any
of
the
serum
chemistry,
hematological,
or
urinary
parameters
measured,
although
several
serum
chemistry
parameters
were
significantly
different
than
controls,
particularly
at
the
high
dose.
The
authors
reported
a
trend
toward
higher
values
for
hemoglobin,
total
red
blood
cell
and
white
blood
cell
counts,
and
fibrinogen
in
all
treated
animals
(
no
data
reported).
No
remarkable
findings
were
observed
at
necropsy
(
gross
observation).

Sporadic
relative
or
absolute
organ
weight
changes
were
reported,
mostly
at
the
high
dose,
but
these
were
not
considered
as
compound
related.
The
only
organ
weight
change
that
showed
a
clear
dose­
dependence
at
multiple
doses
was
for
absolute
liver
weight
in
females,
which
was
significantly
increased
12%
over
controls
(
p

0.05)
beginning
at
23
mg/
kg/
day,
up
to
22%

above
controls
at
90
mg/
kg/
day.
Liver
weights
relative
to
body
weight
and
brain
weight
were
also
affected,
but
only
at
higher
doses.
However,
no
corresponding
significant
elevation
of
serum
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
10
levels
of
hepatic
enzymes
was
observed
in
females,
and
in
the
absence
of
histopathology
data,
it
is
unclear
if
the
increased
liver
weight
in
females
is
an
adverse
response.
The
authors
could
not
identify
specific
target
organs
for
DBAN
toxicity,
and
concluded
that
decreased
body
weight
was
the
most
sensitive
indicator
of
toxicity.
In
males,
no
decrease
in
body
weight
was
observed
at
23
mg/
kg/
day;
body
weight
was
decreased
by
more
than
20%
in
a
dose­
related
manner
at
45
and
90
mg/
kg/
day
(
p

0.05).
No
effect
on
body
weight
was
noted
in
females.
The
authors
stated
that
the
NOAEL
for
DBAN
was
45
mg/
kg/
day,
but
the
decreased
body
weight
in
male
rats
exposed
to
45
mg/
kg/
day
suggests
that
the
NOAEL
in
this
study
is
23
mg/
kg/
day
and
the
LOAEL
is
45
mg/
kg/
day.
The
data
sets
for
body
weight
in
males
and
relative
liver
weight
in
females
were
further
analyzed
to
determine
benchmark
doses
(
BMDs)
according
to
draft
EPA
Guidance
(
U.
S.

EPA,
2000c)
to
identify
alternative
critical
effect
levels.
The
results
of
the
modeling
are
described
in
detail
in
Appendix
A.
A
BMDL
of
16
mg/
kg/
day
for
decreased
body
weight
in
males
was
selected
as
the
most
appropriate
modeling
result
to
serve
as
the
basis
for
the
quantitative
doseresponse
assessment.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
11
Table
V­
2.
Organ
weights
and
ratios
of
CD
rats
exposed
to
dibromoacetonitrile
by
gavage
for
14
days.
a
Parameterb
Vehicle
(
corn
oil)

Male
23
mg/
kg/
day
Male
45
mg/
kg/
day
Male
90
mg/
kg/
day
Male
Vehicle
(
corn
oil)

Female
23
mg/
kg/
day
Female
45
mg/
kg/
day
Female
90
mg/
kg/
day
Female
Body
Weight
275.5
±
3.2
274.6
±
5.5
243.7
±
4.6*
209.4
±
10.1*
176.5
±
6.8
184.9
±
2.4
181.1
±
2.8
165.9
±
5.4
Brain
%
body
weight
1.81
±
0.05
0.66
±
0.02
1.77
±
0.06
0.65
±
0.03
1.76
±
0.03
0.73
±
0.02*
1.66
±
0.05
0.80
±
0.06*
1.67
±
0.04
0.96
±
0.04
1.68
±
0.04
0.91
±
0.03
1.64
±
0.04
0.91
±
0.03
1.69
±
0.06
1.03
±
0.04
Liver
%
body
weight
12.91
±
0.34
4.69
±
0.15
11.60
±
1.24
4.25
±
0.45
11.47
±
1.2
4.71
±
0.49
10.18
±
1.89
4.72
±
0.83
8.59
±
0.26
4.89
±
0.12
9.65
±
0.25*

5.22
±
0.11
9.73
±
0.20*

5.39
±
0.15
10.50
±
0.44*

6.34
±
0.24*

Spleen
%
body
weight
0.64
±
0.03
0.23
±
0.01
0.59
±
0.05
0.21
±
0.02
0.56
±
0.05
0.23
±
0.02
0.42
±
0.05*

0.20
±
0.02
0.48
±
0.03
0.27
±
0.01
0.45
±
0.02
0.24
±
0.01
0.52
±
0.03
0.29
±
0.02
0.42
±
0.04
0.26
±
0.02
Lungs
%
body
weight
1.69
±
0.10
0.61
±
0.03
1.49
±
0.09
0.54
±
0.04
1.44
±
0.07
0.59
±
0.03
1.30
±
0.14*

0.62
±
0.06
1.58
±
0.19
0.88
±
0.08
1.42
±
0.09
0.77
±
0.05
1.25
±
0.06
0.69
±
0.04
1.09
±
0.06*

0.66
±
0.04*

Thymus
%
body
weight
0.51
±
0.03
0.19
±
0.01
0.52
±
0.04
0.19
±
0.01
0.46
±
0.02
0.19
±
0.01
0.24
±
0.06*

0.11
±
0.03*
0.40
±
0.03
0.23
±
0.02
0.44
±
0.03
0.24
±
0.02
0.40
±
0.01
0.22
±
0.01
0.21
±
0.02*

0.13
±
0.01*

Kidneys
%
body
weight
2.50
±
0.10
0.91
±
0.05
2.39
±
0.09
0.87
±
0.03
2.08
±
0.06*

0.85
±
0.02
2.13
±
0.15*

1.02
±
0.06
1.56
±
0.06
0.89
±
0.02
1.64
±
0.04
0.89
±
0.02
1.62
±
0.04
0.89
±
0.03
1.53
±
0.05
0.92
±
0.03
Testes/
Ovaries
%
body
weight
2.70
±
0.09
0.99
±
0.04
2.71
±
0.07
0.99
±
0.03
2.67
±
0.06
1.10
±
0.03
2.69
±
0.06
1.30
±
0.07*
0.11
±
0.01
0.06
±
0.00
0.14
±
0.01
0.07
±
0.01
0.14
±
0.01
0.08
±
0.01
0.12
±
0.02
0.07
±
0.01
Adapted
from
Hayes
et
al.
(
1986).

a
All
data
expressed
as
mean
±
SEM.

b
All
absolute
weights
are
presented
in
grams.

*
Significantly
different
from
vehicle
control
(
p

0.05).
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
12
Table
V­
3.
Serum
Chemistry
Values
for
CD
Rats
Exposed
to
Dibromoacetonitrile
(
DBAN)
for
14
days.
a
Parameter
Vehicle
(
corn
oil)

Male
23
mg/
kg/
day
Male
45
mg/
kg/
day
Male
90
mg/
kg/
day
Male
Vehicle
(
corn
oil)

Female
23
mg/
kg/
day
Female
45mg/
kg/

day
Female
90
mg/
kg/
day
Female
Serum
Glutamate
Pyruvate
Transaminase
(
IU/
L)
168
±
55
75
±
9
76
±
11
77
±
19
74
±
6
62
±
8
74
±
14
68
±
11
Serum
Glutamate
Oxaloacetic
Transaminase
(
IU/
L)
339
±
105
142
±
14
172
±
26
273
±
81
178
±
32
138
±
9
182
±
21
161
±
15
Alkaline
Phosphatase
(
IU/
L)
360
±
41
466
±
33
247
±
29
187
±
8*
236
±
26
299
±
29
260
±
24
179
±
34
5'­
Nucleotidase
(
IU/
L)
18
±
3
17
±
1
13
±
2
14
±
1
25
±
4
22
±
2
28
±
2
18
±
4
Protein
(
g/
dL)
6.3
±
0.1
6.5
±
0.1
6.2
±
0.1
4.7
±
0.4*
6.7
±
0.1
6.2
±
0.3
6.2
±
0.1
5.3
±
0.2*

Albumin
(
g/
dL)
4.0
±
0.1
4.0
±
0.0
4.0
±
0.1
2.9
±
0.3*
4.3
±
0.2
4.0
±
0.1
4.2
±
0.1
3.5
±
0.1*

Globulin
(
g/
dL)
2.3
±
0.1
2.5
±
0.1
2.2
±
0.1
1.8
±
0.1*
2.5
±
0.1
2.2
±
0.2
2.1
±
0.1
1.8
±
0.1*

Alb/
globulin
ratio
1.8
±
0.1
1.6
±
0.1
1.8
±
0.1
1.6
±
0.2
1.7
±
0.1
1.8
±
0.1
2.1
±
0.1
2.0
±
0.1
Glucose
(
mg/
dL)
265
±
47
212
±
41
217
±
40
130
±
1
212
±
17
171
±
18
197
±
22
138
±
17
Cholesterol
(
mg/
dL)
66
±
2
64
±
2
74
±
4
95
±
5*
66
±
4
58
±
6
66
±
6
71
±
3
Bilirubin
(
mg/
dL)
0.3
±
0.0
0.4
±
0.0
0.2
±
0.0
0.3
±
0.1
0.3
±
0.0
0.2
±
0.0
0.3
±
0.0
0.2
±
0.0
BUN
(
mg/
dL)
15
±
1
13
±
1
10
±
1
18
±
6
15
±
2
13
±
2
13
±
1
15
±
2
Creatinine
(
mg/
dL)
1.3
±
0.2
0.8
±
0.1
0.9
±
0.1
0.7
±
0.0*
1.2
±
0.1
1.0
±
0.1
1.0
±
0.1
0.8
±
0.0*

BUN/
creatinine
ratio
13
±
2
16
±
2
12
±
1*
24
±
8
12
±
2
15
±
4
14
±
2
19
±
3
Calcium
(
mg/
dL)
11.3
±
0.4
12.5
±
0.6
11.8
±
0.4
9.6
±
0.6
11.9
±
0.4
11.1
±
0.3
12.1
±
0.5
11.1
±
0.3
Phosphorus
(
mg/
dL)
12.0
±
0.4
11.2
±
0.4
11.3
±
0.5
7.6
±
0.7*
11.6
±
0.1
10.1
±
0.6
10.8
±
0.5
9.3
±
0.7*

Chloride
(
mEq/
L)
99
±
2
99
±
2
99
±
1
100
±
2
98
±
1.4
102
±
0.7
100
±
1.2
101
±
0.7
Adapted
from
Hayes
et
al.
(
1986).

a
All
data
expressed
as
mean
±
SEM.

*
Significantly
different
from
vehicle
control
(
p

0.05).
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
13
Table
V­
4.
Body
Organ
Weights
and
Ratios
of
CD
rats
Exposed
to
Dichloroacetonitrile
by
Gavage
for
14
days.
a
Parameterb
Vehicle
(
corn
oil)

Male
12
mg/
kg/
day
Male
23
mg/
kg/
day
Male
45
mg/
kg/
day
Male
90
mg/
kg/
day
Male
Vehicle
(
corn
oil)

Female
12
mg/
kg/
day
Female
23
mg/
kg/
day
Female
45
mg/
kg/
day
Female
90
mg/
kg/
day
Female
Body
Weight
157
±
5
170
±
7
147
±
5
137
±
5
115
±
4
148
±
6
147
±
6
143
±
4
146
±
5
113
±
5
Brain
%
body
weight
1.68
±
0.09
1.06
±
0.03
1.63
±
0.12
0.99
±
0.09
1.66
±
0.03
1.14
±
0.04
1.56
±
0.04
1.15
±
0.04
1.52
±
0.05*

1.34
±
0.06*
1.61
±
0.04
1.09
±
0.04
1.52
±
0.08
1.04
±
0.04
1.47
±
0.06
1.03
±
0.03
1.61
±
0.05
1.11
±
0.07
1.45
±
0.04
1.30
±
0.06*

Liver
%
body
weight
8.15
±
0.31
5.19
±
0.10
9.96
±
0.58
5.85
±
0.16*
9.71
±
0.64
6.57
±
0.30*
10.11
±
0.52*

7.37
±
0.18*
8.63
±
0.48
7.50
±
0.26*
8.06
±
0.53
5.40
±
.20
8.49
±
0.35
5.81
±
0.15
10.56
±
0.50*

7.37
±
0.27*
11.14
±
0.57*

7.59
±
0.23*
7.78
±
0.92
7.07
±
0.82*

Spleen
%
body
weight
0.54
±
0.04
0.34
±
0.02
0.62
±
0.07
0.36
±
0.04
0.46
±
0.03
0.31
±
0.01
0.47
±
0.05
0.34
±
0.03
0.35
±
0.03*

0.30
±
0.02
0.44
±
0.03
0.30
±
0.02
0.43
±
0.04
0.30
±
0.03
0.40
±
0.02
0.28
±
0.02
0.43
±
0.02
0.29
±
0.02
0.29
±
0.02*

0.26
±
0.01
Lungs
%
body
weight
1.27
±
0.07
0.81
±
0.04
1.37
±
0.11
0.82
±
0.07
1.26
±
0.08
0.85
±
0.05
1.17
±
0.07
0.86
±
0.05
1.02
±
0.09
0.89
±
0.07
1.35
±
0.09
0.91
±
0.06
1.17
±
0.09
0.80
±
0.05
1.13
±
0.07
0.79
±
0.05
1.23
±
0.10
0.84
±
0.06
0.96
±
0.07*

0.86
±
0.07
Thymus
%
body
weight
0.43
±
0.04
0.27
±
0.02
0.55
±
0.05
0.32
±
0.03
0.38
±
0.02*

0.26
±
0.02
0.33
±
0.02*

0.24
±
0.02
0.27
±
0.02*

0.24
±
0.02*
0.48
±
0.03
0.33
±
0.02
0.42
±
0.04
0.29
±
0.03
0.38
±
0.03
0.27
±
0.02
0.42
±
0.02
0.29
±
0.02
0.32
±
0.02*

0.29
±
0.02
Kidneys
%
body
weight
1.53
±
0.05
0.98
±
0.03
1.85
±
0.08*

1.10
±
0.04*
1.52
±
0.05
1.03
±
0.02
1.48
±
0.06
1.08
±
0.02
1.38
±
0.07
1.20
±
0.04*
1.52
±
0.05
1.03
±
0.03
1.42
±
0.06
0.97
±
0.03
1.47
±
0.04
1.03
±
0.03
1.51
±
0.06
1.03
±
0.02
1.28
±
0.05*

1.14
±
0.04
Testes/
Ovaries
%
body
weight
1.54
±
0.13
0.97
±
0.07
1.74
±
0.15
1.00
±
0.06
1.54
±
0.12
1.03
±
0.06
1.26
±
0.13
0.90
±
0.09
1.25
±
0.14
1.07
±
0.11
0.11
±
0.01
0.08
±
0.01
0.10
±
0.01
0.07
±
0.01
0.10
±
0.01
0.07
±
0.01
0.10
±
0.01
0.07
±
0.01
0.08
±
0.01
0.07
±
0.01
Adapted
from
Hayes
et
al.
(
1986).

a
All
data
expressed
as
mean
±
SEM.

b
All
absolute
weights
are
presented
in
grams.

*
Significantly
different
from
vehicle
control
(
p

0.05).
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
14
Table
V­
5.
Serum
Chemistry
Values
for
CD
Rats
Exposed
to
Dichloroacetonitrile
(
DCAN)
for
14
daysa.

Parameter
Vehicle
(
corn
oil)

Male
12
mg/
kg/
day
Male
23
mg/
kg/
day
Male
45
mg/
kg/
day
Male
90
mg/
kg/
day
Male
Vehicle
(
corn
oil)

Female
12
mg/
kg/
day
Female
23
mg/
kg/
day
Female
45
mg/
kg/
day
Female
90
mg/
kg/
day
Female
Serum
Glutamate
Pyruvate
Transaminase
(
IU/
L)
70
±
3
71
±
6
89
±
22
87
±
15
88
±
14
57
±
6
66
±
7
63
±
4
63
±
7
134
±
38*

Serum
Glutamate
Oxaloacetic
Transaminase
(
IU/
L)
167
±
13
266
±
22
213
±
53
276
±
40
233
±
27
205
±
8
198
±
6
201
±
33
181
±
28
340
±
96
Alkaline
Phosphatase
(
IU/
L)
457
±
63
459
±
39
384
±
21
398
±
50
712
±
84*
261
±
24
381
±
48
352
±
45
384
±
35*
651
±
159*

5'­
Nucleotidase
(
IU/
L)
14
±
1
16
±
1
14
±
2
17
±
2
23
±
2*
23
±
3
20
±
2
16
±
1
18
±
1
24
±
2
Protein
(
g/
dL)
5.8
±
0.2
6.0
±
0.2
5.7
±
0.2
5.5
±
0.2
6.0
±
0.1
6.4
±
0.3
6.2
±
0.2
6.4
±
0.3
5.9
±
0.2
6.1
±
0.4
Albumin(
g/
dL)
4.4
±
0.1
4.3
±
0.1
4.3
±
0.1
4.3
±
0.1
4.8
±
0.1
4.6
±
0.1
4.4
±
0.1
4.5
±
0.1
4.7
±
0.1
4.8
±
0.2
Globulin
(
g/
dL)
1.4
±
0.1
1.7
±
0.1
1.4
±
0.2
1.2
±
0.2
1.2
±
0.2
1.8
±
0.2
1.9
±
0.2
1.9
±
0.2
1.2
±
0.1
1.3
±
0.3
Alb/
globulin
ratio
3.1
±
0.1
2.6
±
0.1*
3.4
±
0.5
3.8
±
0.5
4.5
±
0.7
2.8
±
0.4
2.4
±
0.2
2.4
±
0.2
3.9
±
0.3
4.0
±
0.7
Glucose
(
mg/
dL)
154
±
11
129
±
16
138
±
6
134
±
10
112
±
16
191
±
19
162
±
30
142
±
17
160
±
6
137
±
15
Cholesterol
(
mg/
dL)
90
±
5
93
±
2
91
±
6
80
±
10
81
±
11
84
±
3
86
±
4
99
±
6
82
±
8
73
±
8
Bilirubin
(
mg/
dL)
0.2
±
0.0
0.2
±
0.0
0.3
±
0.0
0.4
±
0.0*
0.6
±
0.1
0.2
±
0.0
0.2
±
0.0
0.2
±
0.0
0.4
±
0.0*
0.7
±
0.1*

BUN
(
mg/
dL)
17
±
2
17
±
2
15
±
2
17
±
2
15
±
1
17
±
1
17
±
1
15
±
2
20
±
3
23
±
2
Creatinine
(
mg/
dL)
1.7
±
0.2
1.4
±
0.2
1.4
±
0.1
1.5
±
0.1
1.4
±
0.1
1.6
±
0.2
1.3
±
0.1
1.2
±
0.1
1.6
±
0.1
1.6
±
0.2
BUN/
creatinine
ratio
11
±
3
12
±
1
11
±
1
12
±
2
11
±
1
11
±
1
13
±
1
13
±
2
12
±
1
14
±
1
Calcium
(
mg/
dL)
11.4
±
0.1
11.0
±
0.2
11.3
±
0.2
9.8
±
1.5
11.7
±
0.3
11.7
±
0.5
11.1
±
0.3
11.3
±
0.2
11.9
±
0.3
11.8
±
0.5
Phosphorus
(
mg/
dL)
12.3
±
0.6
12.0
±
0.8
12.5
±
0.6
11.7
±
0.5
12.9
±
0.4
11.4
±
0.9
10.9
±
0.5
11.4
±
0.9
12.7
±
0.1
9.9
±
0.4
Chloride(
mEq/
L)
100
±
1
100
±
1
101
±
1
101
±
2
101
±
1
100
±
1
99
±
1
100
±
1
101
±
1
101
±
2
Adapted
from
Hayes
et
al.
(
1986).
aAll
data
expressed
as
mean
±
SEM
*
Significantly
different
from
vehicle
control
at
(
p

0.05)
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
15
EPA/
OW/
OST/
HECD
Final
Draft
In
a
similar
14­
day
repeated
dosing
study
with
DCAN,
Hayes
et
al.
(
1986)
administered
gavage
doses
of
0,
12,
23,
45,
or
90
mg/
kg/
day
in
corn
oil
to
CD
rats
(
10
animals/
sex/
dose).

Food
and
water
consumption
rates
were
not
measured.
No
mortality
was
reported
for
any
treatment
group.
In
males,
depression
in
body
weight
to
94%,
87%,
and
73%
of
body
weight
in
control
animals
was
observed
at
23,
45,
and
90
mg/
kg/
day,
respectively,
as
shown
in
Table
V­
4.

In
females,
body
weight
was
decreased
to
76%
of
controls
at
90
mg/
kg/
day.
Several
serum
markers
for
organ
toxicity
were
increased
in
treated
animals
as
shown
in
Table
V­
5.
Significantly
increased
levels
of
serum
glutamic
pyruvate
transaminase
(
SGPT)
in
females
at
90
mg/
kg/
day,
and
alkaline
phosphatase
(
ALP)
levels
at
90
mg/
kg/
day
in
males
and
at
45
and
90
mg/
kg/
day
in
females
were
reported.
Although
the
authors
did
not
consider
these
changes
to
be
compoundrelated
adverse
effects,
they
did
not
provide
a
reason.
We
considered
these
changes
to
be
adverse,
based
on
the
magnitude
of
the
change,
and
the
supporting
data
for
DCAN
in
females
in
the
subchronic
study.
No
other
consistent
dose­
dependent
adverse
effects
were
observed
in
any
of
the
other
serum
chemistry,
hematological,
or
urinary
parameters
measured.
The
authors
reported
a
trend
toward
elevated
red
blood
cell
and
white
blood
cell
counts
in
all
treated
animals,

but
no
data
were
provided.
No
remarkable
findings
were
observed
at
necropsy
(
gross
observation).
Relative
liver
weight
was
significantly
increased
(
p

0.05)
in
male
and
female
rats.

The
relative
liver
weights
in
males
were
13%,
26%,
42%,
and
45%
greater
than
controls
at
doses
of
12,
23,
45,
and
90
mg/
kg/
day,
respectively.
Absolute
liver
weight,
in
contrast,
was
significantly
increased
(
p

0.05)
only
in
the
45
mg/
kg/
day
dose
group.
In
female
rats,
both
relative
and
absolute
liver
weights
were
elevated
beginning
at
23
mg/
kg/
day,
with
relative
liver
weights
36%,
40%,
and
31%
greater
than
controls
at
23,
45,
and
90
mg/
kg/
day,
respectively.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
16
EPA/
OW/
OST/
HECD
Final
Draft
Liver
weight
changes
would
be
considered
as
potentially
adaptive
in
the
absence
of
other
signs
of
hepatic
injury.
However,
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.
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.
The
data
sets
for
body
weight
and
relative
liver
weight
in
males
and
females
were
further
analyzed
to
determine
benchmark
doses
(
BMDs)
according
to
draft
EPA
Guidance
(
U.
S.
EPA,
2000c)
to
identify
alternative
critical
effect
levels.
The
results
of
the
modeling
are
described
in
detail
in
Appendix
A.

A
BMDL
of
5
mg/
kg/
day
for
relative
liver
weight
in
males
was
selected
as
the
most
appropriate
modeling
result
to
serve
as
the
basis
for
the
quantitative
dose­
response
assessment.

B.
Long­
Term
Exposures
No
studies
were
identified
that
evaluated
the
toxicity
of
long­
term
exposure
to
BCAN
or
TCAN
by
any
route.
No
studies
were
identified
that
evaluated
the
toxicity
of
long­
term
exposure
to
DBAN
or
DCAN
by
the
inhalation
or
dermal
routes.

The
subchronic
toxicity
of
DBAN
has
been
evaluated
by
the
NTP
(
2002)
in
B6C3F1
mice
and
F344
rats
as
part
of
initial
dose­
range
finding
studies
for
chronic
exposure
studies
that
are
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
17
EPA/
OW/
OST/
HECD
Final
Draft
currently
in
progress.
For
the
mouse
study,
DBAN
was
administered
in
drinking
water
for
13
weeks
to
male
and
female
B6C3F1
mice
(
10/
sex/
dose)
at
concentrations
of
0,
12.5,
25,
50,
100,

and
200
mg/
L.
The
corresponding
doses
reported
by
the
study
authors
were
0,
1.6,
3.2,
5.6,

10.7,
and
17.9
mg/
kg/
day
for
males
and
0,
1.6,
3.0,
6.1,
11.1,
and
17.9
mg/
kg/
day
for
females.

Animals
were
observed
for
clinical
signs
of
toxicity,
as
well
as
body
weight,
organ
weight
and
pathology,
hematology,
and
clinical
chemistry.
A
separate
set
of
animals
(
10/
sex/
dose)
were
exposed
to
the
same
concentrations
as
the
main
study
groups
for
26
days,
but
were
co­
exposed
to
DBAN
and
5­
bromo­
2­
deoxyuridine
during
the
last
five
days
of
this
period.
These
animals
were
used
to
collect
tissue
sample
for
analysis
of
induce
cell
proliferation.
Decreased
water
consumption
and
decreased
body
weight
were
the
only
effects
related
to
DBAN
treatment.

Decreased
water
consumption
was
observed
in
both
males
and
females
at
DBAN
concentrations
of
50
mg/
L
and
higher.
A
slight
and
transient
decrease
in
body
weight
gain
was
observed;

terminal
body
weights
were
94%
of
controls
in
high­
dose
males
and
96%
of
controls
in
high
dose
females.
These
small
changes
are
not
judged
as
toxicologically­
significant.
Based
on
the
minimal
effects
observed
for
DBAN
in
this
study,
the
NOAEL
is
17.9
mg/
kg/
day
for
males
and
females.

No
LOAEL
was
identified.

For
the
rat
study,
DBAN
was
administered
in
drinking
water
for
13
weeks
to
male
and
female
F344
rats
(
10/
sex/
dose)
at
concentrations
of
0,
12.5,
25,
50,
100,
and
200
mg/
L.
The
corresponding
doses
reported
by
the
study
authors
were
0,
0.9,
1.8,
3.3,
6.2,
and
11.3
mg/
kg/
day
for
males
and
0,
1.0,
1.9,
3.8,
6.8,
and
12.6
mg/
kg/
day
for
females.
Animals
were
observed
for
clinical
signs
of
toxicity,
as
well
as
body
weight,
organ
weight
and
pathology,
hematology,
and
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
18
EPA/
OW/
OST/
HECD
Final
Draft
clinical
chemistry.
A
separate
set
of
animals
(
10/
sex/
dose)
were
exposed
to
the
same
concentrations
as
the
main
study
groups
for
26
days,
but
were
co­
exposed
to
DBAN
and
5­

bromo­
2­
deoxyuridine
during
the
last
five
days
of
this
period.
These
animals
were
used
to
collect
tissue
samples
for
analysis
of
cell
proliferation.
Decreased
water
consumption
and
decreased
body
weight
were
the
only
effects
related
to
DBAN
treatment.
Slight
changes
in
clinical
chemistry
and
hematology
findings
were
considered
by
the
study
authors
to
be
related
to
decreased
water
consumption.
Decreased
water
consumption
was
observed
in
males
at
DBAN
concentrations
of
50
mg/
L
and
higher
and
in
females
at
the
two
highest
concentrations.
A
slight
decrease
in
body
weight
gain
was
observed
for
high
dose
males
and
females.
Terminal
body
weights
were
94%
of
controls
in
high­
dose
males
and
95%
of
controls
in
high
dose
females.
These
small
changes
are
not
judged
as
toxicologically­
significant.
Based
on
the
minimal
effects
observed
for
DBAN
in
this
study,
the
NOAEL
is
11.3
mg/
kg/
day
for
males
and
12.6
mg/
kg/
day
for
females.
No
LOAEL
was
identified.

In
a
90­
day
study
of
DBAN
toxicity,
Hayes
et
al.
(
1986)
administered
doses
of
0,
6,
23,

or
45
mg/
kg/
day
in
corn
oil
by
gavage
to
groups
of
CD
rats
(
20
animals/
sex/
dose).
No
compound­
related
deaths
occurred
during
this
study.
Body
weights
were
depressed
to
79%
of
controls
at
the
end
of
the
study
in
males,
but
not
in
females,
at
the
highest
dose
tested
(
45
mg/
kg/

day).
Food
and
water
consumption
rates
were
not
measured.
Observed
changes
in
the
serum
chemistry,
hematological,
urinary
parameters,
and
organ
weight
were
generally
not
dose­
related
and
were
not
considered
to
be
compound­
related
(
Tables
V­
6
and
V­
7).
The
only
exceptions
were
significantly
increased
ALP
in
females
at
45
mg/
kg/
day
and
a
significant
increase
in
relative
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
19
EPA/
OW/
OST/
HECD
Final
Draft
liver
weight
(
but
not
absolute
weight)
in
males
at
45
mg/
kg/
day.
Interim
serum
biochemistry
analyses
at
one
and
two
months
of
exposure
also
revealed
no
treatment­
related
effects.
No
remarkable
findings
were
apparent
at
gross
necropsy
and
the
authors
did
not
identify
a
specific
target
tissue
for
DBAN.
The
finding
that
effects
on
the
liver
in
this
subchronic
study
were
comparable
to
the
effects
following
the
14­
day
exposure
at
the
same
doses
suggests
that
the
liver
weight
changes
seen
in
the
14­
day
study
were
adaptive,
or
at
least
that
tolerance
developed
to
the
repeated
dosing.
In
addition,
no
clear
increase
in
serum
enzyme
markers
of
hepatic
injury
were
observed
in
males,
even
though
they
had
a
greater
increase
in
relative
liver
weight
than
females.

Based
on
this
discordance
between
serum
biochemistry
and
liver
weight
findings,
and
in
light
of
the
results
from
the
14­
day
study,
the
observed
liver
weight
changes
were
not
judged
sufficiently
adverse
to
serve
as
the
basis
for
the
dose­
response
assessment.
Based
on
decreased
body
weight
in
males
as
the
most
sensitive
endpoint,
the
NOAEL
for
this
study
is
23
mg/
kg/
day
and
the
LOAEL
is
45
mg/
kg/
day.

The
body
weight
data
in
males
were
further
analyzed
to
determine
benchmark
doses
(
BMDs)
according
to
draft
EPA
Guidance
(
U.
S.
EPA,
2000c)
to
identify
alternative
critical
effect
levels.
The
results
of
the
modeling
are
described
in
detail
in
Appendix
A.
A
BMDL
of
20
mg/
kg/
day
for
decreased
body
weight
in
males
was
selected
as
the
most
appropriate
modeling
result
to
serve
as
the
basis
for
the
quantitative
dose­
response
assessment.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
20
EPA/
OW/
OST/
HECD
Final
Draft
Table
V­
6.
Body
and
Organ
Weights
for
CD
Rats
Exposed
to
Dibromoacetonitrile
(
DBAN)
by
Gavage
for
90
Daysa.

Parameterb
Vehicle
(
corn
oil)
Male
6
mg/
kg/
day
Male
23
mg/
kg
day
Male
45
mg/
kg/
day
Male
Vehicle
(
corn
oil)
Female
6
mg/
kg/
day
Female
23
mg/
kg/
day
Female
45
mg/
kg/
day
Female
Body
Weight
556.0
±
12.7
545.9
±
13.5
523.1
±
11.4
438.6
±
7.5*
292.9
±
5.8
279.6
±
7.3
291.7
±
9.5
274.9
±
7.2
Brain
%
body
weight
1.96
±
0.04
0.36
±
0.01
1.98
±
0.05
0.37
±
0.01
1.95
±
0.04
0.38
±
0.01
1.93
±
0.4
0.44
±
0.01*
1.83
±
0.03
0.63
±
0.02
1.81
±
0.04
0.65
±
0.02
1.76
±
0.04
0.61
±
0.01
1.76
±
0.04
0.64
±
0.02
Liver
%
body
weight
21.85
±
0.99
3.93
±
0.15
21.70
±
0.68
3.98
±
0.08
22.39
±
0.84
4.27
±
0.12
19.7
±
0.67
4.44
±
0.13*
11.62
±
0.37
3.98
±
0.12
11.1
±
0.36
3.98
±
0.10
12.31
±
0.33
4.25
±
0.11
11.95
±
0.39
4.36
±
0.12
Spleen
%
body
weight
0.84
±
0.04
0.15
±
0.01
0.84
±
0.03
0.15
±
0.01
0.78
±
0.04
0.15
±
0.01
0.72
±
0.04
0.16
±
0.01
0.51
±
0.03
0.17
±
0.01
0.52
±
0.02
0.19
±
0.01
0.55
±
0.03
0.19
±
0.01
0.57
±
0.04
0.21
±
0.01
Lungs
%
body
weight
3.22
±
0.14
0.58
±
0.03
2.95
±
0.09
0.54
±
0.02
3.09
±
0.11
0.59
±
0.02
2.73
±
0.09*
0.63
±
0.03
2.27
±
0.15
0.78
±
0.05
2.34
±
0.16
0.85
±
0.07
2.35
±
0.14
0.81
±
0.05
2.47
±
0.17
0.90
±
0.06
Thymus
%
body
weight
0.55
±
0.03
0.10
±
0.01
0.50
±
0.03
0.09
±
0.01
0.53
±
0.02
0.10
±
0.00
0.43
±
0.03*
0.10
±
0.01
0.40
±
0.03
0.14
±
0.01
0.40
±
0.03
0.15
±
0.01
0.34
±
0.02
0.12
±
0.01
0.26
±
0.03*
1.10
±
0.01*

Kidneys
%
body
weight
3.67
±
0.10
0.66
±
0.02
3.56
±
0.09
0.66
±
0.02
3.62
±
0.10
0.69
±
0.01
3.16
±
0.10*
0.72
±
0.02*
2.10
±
0.05
0.72
±
0.02
1.98
±
0.06
0.71
±
0.02
2.01
±
0.70
0.70
±
0.02
1.98
±
0.07
0.72
±
0.02
Testes/
Ovaries
%
body
weight
3.58
±
0.05
0.65
±
0.01
3.76
±
0.07
0.69
±
0.02
3.49
±
0.60
0.67
±
0.02
3.51
±
0.15
0.80
±
0.03*
0.17
±
0.01
0.06
±
0.00
0.17
±
0.01
0.06
±
0.00
0.16
±
0.01
0.05
±
0.00
0.17
±
0.01
0.06
±
0.60
Adapted
from
Hayes
et
al.
(
1986).

aAll
data
expressed
as
mean
±
SEM
bAll
absolute
weights
are
presented
in
grams.
*
Significantly
different
from
vehicle
control
(
p

0.05).
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
21
Table
V­
7.
Serum
Chemistry
Values
for
CD
Rats
Exposed
to
Dibromoacetonitrile
(
DBAN)
by
Gavage
for
90
days.
a
Parameter
Vehicle
(
corn
oil)

Male
6
mg/
kg/
day
Male
23
mg/
kg/
day
Male
45
mg/
kg/
day
Male
Vehicle
(
corn
oil)

Female
6
mg/
kg/
day
Female
23
mg/
kg/
day
Female
45
mg/
kg/
day
Female
Serum
Glutamate
Pyruvate
Transaminase
(
IU/
L)
44
±
4
69
±
3
42
±
4
44
±
5
46
±
3
35
±
3
37
±
3
32
±
3*

Serum
Glutamate
Oxaloacetic
Transaminase
(
IU/
L)
176
±
12
245
±
52
190
±
30
181
±
13
207
±
31
169
±
31
174
±
10
161
±
11
Alkaline
Phosphatase
(
IU/
L)
191
±
16
171
±
28
192
±
23
157
±
16
134
±
11
130
±
9
138
±
19
208
±
32*

5'­
Nucleotidase
(
IU/
L)
15
±
1
17
±
2
16
±
1
13
±
1
26
±
2
26
±
2
24
±
2
19
±
1*

Protein
(
g/
dL)
8.2
±
0.4
8.2
±
0.3
8.1
±
0.2
7.0
±
0.2*
7.3
±
0.2
7.5
±
0.2
7.1
±
0.2
6.9
±
0.2
Albumin(
g/
dL)
5.2
±
0.1
5.3
±
0.2
5.8
±
0.1*
5.8
±
0.1*
6.1
±
0.1
6.6
±
0.1*
6.1
±
0.1
5.6
±
0.1*

Globulin
(
g/
dL)
2.9
±
0.5
2.9
±
0.4
2.3
±
0.2
1.2
±
0.1*
1.2
±
0.1
0.9
±
0.2
1.1
±
0.1
1.3
±
0.2
Alb/
globulin
ratio
2.5
±
0.6
2.3
±
0.3
2.8
±
0.4
5.8
±
0.8
5.7
±
0.6
9.0
±
1.3*
6.3
±
0.8
5.3
±
0.8
Glucose
(
mg/
dL)
148
±
7
139
±
7
143
±
8
134
±
5
145
±
8
147
±
5
135
±
10
130
±
7
Cholesterol
(
mg/
dL)
73
±
5
70
±
5
69
±
4
77
±
5
80
±
4
80
±
6
78
±
4
64
±
3*

Bilirubin
(
mg/
dL)
0.6
±
0.1
0.9
±
0.1
0.8
±
0.1
0.7
±
0.0
0.5
±
0.0
0.5
±
0.0
0.4
±
0.1
0.7
±
0.0*

BUN
(
mg/
dL)
17
±
1
15
±
1
17
±
1
18
±
1
18
±
1
18
±
1
18
±
1
15
±
1
Creatinine
(
mg/
dL)
1.0
±
0.1
0.9
±
0.1
1.3
±
0.0
1.0
±
0.0
1.2
±
0.0
1.3
±
0.1
1.0
±
0.1
1.0
±
0.0*

BUN/
creatinine
ratio
17
±
1
17
±
2
14
±
0
19
±
1
15
±
1
14
±
1
18
±
2
15
±
1
Calcium
(
mg/
dL)
11.3
±
0.5
11.7
±
0.2
12.8
±
0.2*
10.9
±
0.2
10.9
±
0.2
11.4
±
0.2
11.1
±
0.2
10.7
±
0.3
Phosphorus
(
mg/
dL)
6.4
±
0.4
6.9
±
0.4
6.7
±
0.2
6.5
±
0.2
4.6
±
0.2
4.8
±
0.1
5.7
±
0.2*
6.2
±
0.3*

Chloride
(
mEq/
L)
102
±
1
101
±
1
102
±
1
102
±
1
102
±
1
101
±
1
102
±
1
101
±
1
Adapted
from
Hayes
et
al.
(
1986).

a
All
data
expressed
as
mean
SEM
*
Significantly
different
from
vehicle
control
(
p

0.05).
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
22
EPA/
OW/
OST/
HECD
Final
Draft
In
a
90­
day
study
of
DCAN
toxicity,
Hayes
et
al.
(
1986)
administered
doses
of
0,
8,
33,

or
65
mg/
kg/
day
by
gavage
in
corn
oil
to
groups
of
CD
rats
(
20
animals/
sex/
dose).
Food
and
water
consumption
data
were
not
reported.
At
65mg/
kg/
day,
50%
of
males
and
25%
of
females
had
died
by
the
completion
of
the
study;
at
33
mg/
kg/
day,
10%
of
males
and
5%
of
females
had
died;
and
at
8
mg/
kg/
day,
5%
of
males
had
died.
Supplementary
information
provided
by
the
authors
indicated
that
there
were
one
to
three
deaths
(
5%
to
15%)
in
the
control
groups,
and
that
some
of
the
deaths
in
the
high­
dose
groups
were
due
to
gavage
error.
Most
of
the
deaths
that
were
judged
to
be
compound­
related
occurred
in
weeks
9
to
10.
Body
weight
was
significantly
depressed
in
male
and
female
rats
at
65
mg/
kg/
day
(
to
73%
of
controls)
and
in
males
at
33
mg/
kg/
day
(
to
81%
of
controls).
Most
of
the
observed
changes
in
the
serum
chemistry,

hematological,
and
urinary
parameters
did
not
appear
to
be
compound­
related
(
Tables
V­
8
and
V­

9).
The
exception
was
alkaline
phosphatase,
which
was
significantly
increased
in
males
and
females
at
the
high
dose,
and
in
males
also
at
33
mg/
kg/
day.
Sporadic
organ
weight
changes
were
observed,
mostly
at
the
high
dose.
Of
these,
a
dose­
dependent
increase
was
seen
only
for
relative
liver
weights.
Relative
liver
weight
(
relative
to
body
weight)
was
significantly
increased
(
p

0.05)

in
males
beginning
at
33
mg/
kg/
day
(
60%
increase),
and
in
females
beginning
at
8
mg/
kg/
day
(
17%
increase).
The
relative
liver
weight
was
also
increased
in
males
(
by
12%)
at
8
mg/
kg/
day.

Although
the
12%
increase
in
relative
liver
weight
in
males
was
not
statistically
significant,
it
is
judged
to
be
biologically
significant,
based
on
the
magnitude
of
the
change
and
the
observation
in
the
14­
day
study
that
males
were
more
sensitive
to
liver
weight
changes
than
females.
Liver
weight
changes
would
be
considered
as
potentially
adaptive
in
the
absence
of
other
signs
of
hepatic
injury.
The
observed
increase
in
serum
levels
of
ALP
activity
gives
greater
weight
to
the
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
23
EPA/
OW/
OST/
HECD
Final
Draft
potential
toxicological
significance
of
the
liver
weight
changes,
although
ALP
is
not
a
liverspecific
enzyme,
and
no
corresponding
increases
in
the
liver­
specific
enzymes
SGPT
or
serum
glutamate
oxaloacetic
transaminase
(
SGOT)
were
observed
at
study
termination
in
the
subchronic
study.
However,
the
toxicological
relevance
of
the
ALP
results
following
subchronic
dosing
is
supported
by
the
increase
in
both
ALP
and
SGPT
observed
in
the
14­
day
study.
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.

The
body
weight
and
relative
liver
weight
data
in
males
and
females
were
further
analyzed
to
determine
benchmark
doses
(
BMDs)
for
these
endpoints
according
to
draft
EPA
Guidance
(
U.
S.
EPA,
2000c)
to
identify
alternative
critical
effect
levels.
The
results
of
the
modeling
are
described
in
detail
in
Appendix
A.
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.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
24
Table
V­
8.
Body
and
Organ
Weights
for
CD
Rats
Exposed
to
Dichloroacetonitrile
(
DCAN)
by
Gavage
for
90
daysa.

Parameterb
Vehicle
(
corn
oil)

Male
8
mg/
kg/
day
Male
33
mg/
kg/
day
Male
65
mg/
kg/
day
Male
Vehicle
(
corn
oil)

Female
8
mg/
kg/
day
Female
33
mg/
kg
day
Female
65
mg/
kg/
day
Female
Body
Weight
541.4
±
13.1
572.1
±
14.5
438.0
±
16.6*
285.6
±
20.6*
309.0
±
6.91
282.5
±
17.4
280.6
±
10.7
225.5
±
8.9*

Brain
%
body
weight
1.86
±
0.06
0.35
±
0.01
1.84
±
0.07
0.33
±
0.02
1.86
±
0.06
0.44
±
0.02*
1.77
±
0.06
0.63
±
0.03*
1.68
±
0.07
0.55
±
0.02
1.69
±
0.07
0.57
±
0.02
1.58
±
0.06
0.58
±
0.03
1.46
±
0.08
0.66
±
0.05
Liver
%
body
weight
21.43
±
0.73
4.0
±
0.10
25.83
±
0.104
4.5
±
0.13
27.26
±
1.06
6.4
±
0.43*
17.91
±
1.60
6.4
±
0.92*
12.10
±
0.42
4.0
±
0.14
14.13
±
0.87
4.7
±
0.29*
16.83
±
0.58*

6.1
±
0.22*
13.88
±
0.705
6.1
±
0.15*

Spleen
%
body
weight
0.72
±
0.03
0.13
±
0.03
0.73
±
0.03
0.13
±
0.02
0.63
±
0.03
0.15
±
0.04
0.50
±
0.06*

0.18
±
0.03
0.54
±
0.02
0.18
±
0.01
0.56
±
0.03
0.19
±
0.01
0.50
±
0.02
0.18
±
0.01
0.450
±
0.020
0.20
±
0.01
Lungs
%
body
weight
2.98
±
0.13
0.56
±
0.02
2.77
±
0.13
0.49
±
0.02*
2.63
±
0.10
0.61
±
0.02
1.83
±
0.06*

0.65
±
0.04
2.04
±
0.10
0.67
±
0.03
2.20
±
0.15
0.74
±
0.05
2.22
±
0.16
0.79
±
0.03*
1.83
±
0.10
0.82
±
0.05*

Thymus
%
body
weight
0.60
±
0.03
0.12
±
0.01
0.79
±
0.04*

0.14
±
0.01
0.62
±
0.03
0.15
±
0.01*
0.31
±
0.03*

0.11
±
0.02
0.50
±
0.03
0.16
±
0.01
0.54
±
0.03
0.18
±
0.01
0.42
±
0.02*

0.15
±
0.01
0.375
±
0.02
0.17
±
0.01
Kidneys
%
body
weight
3.71
±
0.07
0.69
±
0.01
3.88
±
0.12
0.68
±
0.02
3.45
±
0.12
0.81
±
0.04*
2.77
±
0.19*

1.0
±
0.12*
2.23
±
0.06
0.72
±
0.01
2.38
±
0.06
0.80
±
0.02
2.20
±
0.08
0.80
±
0.03
2.25
±
0.12
1.00
±
0.02*

Testes/
Ovaries
%
body
weight
3.42
±
0.08
0.63
±
0.02
3.66
±
0.06
0.65
±
0.02
3.40
±
0.07
0.80
±
0.04*
2.91
±
0.15*

1.0
±
0.12*
0.16
±
0.01
0.05
±
0.004
0.14
±
0.01
0.05
±
0.002
0.14
±
0.01
0.05
±
0.003
0.10
±
0.02*

0.05
±
0.01
Adapted
from
Hayes
et
al.
(
1986).

All
data
expressed
as
mean
±
SEM.

b
All
absolute
weights
are
presented
in
grams.

*
Significantly
different
from
vehicle
control
(
p

0.05).
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
25
Table
V­
9.
Serum
Chemistry
Values
for
CD
Rats
Exposed
to
Dichloroacetonitrile
(
DCAN)
for
90
days.
a
Parameter
Vehicle
(
corn
oil)

Male
8
mg/
kg/
day
Male
33
mg/
kg/
day
Male
65
mg/
kg/
day
Male
Vehicle
(
corn
oil)

Female
8
mg/
kg/
day
Female
33
mg/
kg/
day
Female
65
mg/
kg/
day
Female
Serum
Glutamate
Pyruvate
Transaminase
(
IU/
L)
53
±
1
45
±
2
116
±
60
53
±
4
51
±
3
32
±
2*
27
±
2*
35
±
3*

Serum
Glutamate
Oxaloacetic
Transaminase
(
IU/
L)
195
±
28
140
±
13
265
±
96
163
±
18
120
±
14
127
±
15
124
±
14
126
±
16
Alkaline
Phosphatase
(
IU/
L)
222
±
20
320
±
31
471
±
46*
603
±
79*
228
±
29
226
±
25
286
±
30
499
±
46*

5'­
Nucleotidase
(
IU/
L)
19
±
2
14
±
1
22
±
3
21
±
1
30
±
2
19
±
1*
16
±
1*
17
±
1*

Protein
(
g/
dL)
7.5
±
0.1
7.2
±
0.1
7.0
±
0.1*
6.5
±
0.1
7.1
±
0.1
7.3
±
0.1
7.1
±
0.1
6.6
±
0.1*

Albumin
(
g/
dL)
5.6
±
0.1
6.0
±
0.1*
5.6
±
0.1*
5.4
±
0.1
6.2
±
0.2
6.0
±
0.1
6.0
±
0.1
5.5
±
0.1*

Globulin
(
g/
dL)
1.9
±
0.1
1.2
±
0.1
1.4
±
0.1
1.0
±
0.1
0.9
±
0.1
1.3
±
0.1
1.1
±
0.1
1.1
±
0.1
Alb/
globulin
ratio
3.1
±
0.4
5.3
±
0.4
4.0
±
0.4
5.0
±
1
7.0
±
1
5.1
±
0.4
6.0
±
0.4
5.0
±
1
Glucose
(
mg/
dL)
145
±
4
143
±
3
147
±
4
109
±
4*
150
±
6
127
±
5*
140
±
5
120
±
6*

Cholesterol
(
mg/
dL)
81
±
6
92
±
6
71
±
5
55
±
4*
73
±
3
96
±
6*
80
±
7
54
±
4*

Bilirubin
(
mg/
dL)
0.5
±
0.0
0.5
±
0.0
0.5
±
0.1
0.6
±
0.1
0.5
±
0.0
0.4
±
0.0
0.4
±
0.0
0.5
±
0.0
BUN
(
mg/
dL)
14
±
1
13
±
1
14
±
1
16
±
2
15
±
1
16
±
1
15
±
1
16
±
1
Creatinine
(
mg/
dL)
1.1
±
0.0
1.1
±
0.0
1.2
±
0.0
1.2
±
0.0
1.1
±
0.0
1.2
±
0.1
1.2
±
0.1
1.4
±
0.1*

BUN/
creatinine
ratio
13.8
±
1
11.7
±
0.7
11.4
±
0.7
13.2
±
1
14
±
1
13
±
1
12
±
1
12
±
1
Calcium
(
mg/
dL)
10.1
±
0.2
9.4
±
0.1*
9.6
±
0.2
9.2
±
0.3*
9.7
±
0.1
10.1
±
0.2
9.8
±
0.2
9.8
±
0.2
Phosphorus
(
mg/
dL)
5.7
±
0.3
5.1
±
0.2
6.1
±
0.4
5.9
±
0.3
5.0
±
0.2
5.1
±
0.2
5.4
±
0.2
5.0
±
0.2
Chloride
(
mEq/
L)
105
±
1
103
±
1
103
±
1
107
±
1
107
±
1
106
±
2
105
±
1
106
±
1
Adapted
from
Hayes
et
al.
(
1986).

a.
All
data
expressed
as
mean
±
SEM.

*
Significantly
different
from
vehicle
control
at
p

0.05.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
26
EPA/
OW/
OST/
HECD
Final
Draft
C.
Reproductive
and
Developmental
Effects
No
multigeneration
studies
were
located
on
the
reproductive
effects
of
any
of
the
haloacetonitriles
(
HANs).
No
reproductive
or
developmental
studies
were
identified
for
any
of
the
HANs
following
inhalation
or
dermal
exposure.

Meier
et
al.
(
1985)
evaluated
the
in
vivo
genotoxicity
of
several
drinking
water
disinfectants
and
their
by­
products
in
mice.
As
part
of
this
study,
the
ability
of
a
series
of
disinfectants
to
induce
sperm
head
shape
abnormalities
was
examined
as
a
measure
of
germ
cell
mutagenicity.
Of
the
disinfectants
tested,
only
hypochlorite
induced
a
dose­
related
increase
in
the
percent
of
abnormal
sperm
heads.
To
investigate
whether
this
positive
finding
with
hypochlorite
might
be
due
to
the
in
vivo
formation
of
haloacetonitriles
from
hypochlorite,
groups
of
8
to11­

week
old
male
B6C3F
1
mice
(
10/
dose
group)
were
administered
0,
12.5,
25,
or
50
mg/
kg/
day
of
BCAN,
DBAN,
DCAN,
or
TCAN
in
water
by
gavage
for
5
days.
The
authors
indicated
that
the
highest
total
dose
of
250
mg/
kg
(
5
days
x
50
mg/
kg/
day)
was
selected
to
approximate
the
reported
LD
50
values
for
these
compounds.
Positive
controls
received
five
daily
doses
of
200
mg/
kg
ethylmethanesulfonate
administered
intraperitoneally.
Control
and
test
animals
were
sacrificed
three
or
five
weeks
after
the
last
treatment
and
sperm
recovered
from
the
caudae
epididymides
were
examined
for
abnormal
sperm­
head
morphology.
The
study
authors
reported
no
effects
on
sperm
head
shape
abnormalities
for
any
of
the
HANs
at
doses
up
to
50
mg/
kg/
day.

The
positive
control
yielded
an
increase
in
sperm
head
abnormalities.
The
NOAEL
for
effects
on
sperm
head
abnormalities
in
this
study
is
50
mg/
kg/
day
for
all
of
the
HANs
that
were
tested.
The
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
27
EPA/
OW/
OST/
HECD
Final
Draft
HANs
were
not
tested
for
their
potential
to
cause
micronuclei
or
chromosome
aberrations,
since
no
positive
results
were
seen
for
these
end
points
with
the
initial
disinfectants.

R.
O.
W.
Sciences
(
1997)
conducted
a
reproductive
and
developmental
toxicity
screening
study
for
DBAN.
Based
on
the
results
of
the
dose­
range
finding
studies
(
described
in
detail
in
the
section
on
shorter­
term
toxicity
studies),
the
main
reproductive
and
developmental
toxicity
study
included
concentrations
of
0,
15,
50,
or
150
ppm
DBAN
in
the
drinking
water
of
Sprague­
Dawley
rats
(
10
animals/
dose).
Male
rats,
11
weeks
of
age
on
study
day
1,
were
given
treated
water
on
study
days
6
through
34
or
35
(
the
day
of
necropsy).
The
estimated
doses
resulting
from
exposure
to
0,
15,
50,
or
150
ppm
DBAN
were
0,
1.4,
3.3,
and
8.2
mg/
kg/
day
(
calculated
by
the
study
authors
from
body
weight
and
water
consumption
data).
Males
were
examined
for
clinical
signs
of
toxicity
and
body
weight
at
various
intervals
during
the
study.
At
study
termination,

clinical
pathology
(
hematology
and
clinical
chemistry),
body
and
organ
(
liver,
right
kidney,
spleen,

thymus,
right
testis,
right
epididymis,
right
cauda
epididymis)
weight
measurements,
and
gross
and
histopathology
(
of
the
same
array
of
organs
for
which
organ
weight
was
determined)
were
conducted.
Sperm
analyses
for
control
and
high­
dose
males
included
sperm
motility,
spermatid
head
counts,
sperm
morphology,
and
chromatin
structure.
Sperm
density
was
assessed
for
all
necropsied
males.
All
parameters
evaluated
were
within
normal
limits,
unless
described
below.

The
only
concentration­
related
effects
in
males
were
a
slight
(
4%)
body
weight
decrease
at
150
ppm
that
was
not
statistically
significant,
and
statistically
significant
decreases
in
water
consumption
at
50
ppm
and
150
ppm.
Decreased
water
consumption
(
reported
on
the
basis
of
grams/
kg
body
weight/
day)
ranged
from
68%
to
78%
for
study
days
8,
10,
21,
and
33
in
the
50
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
28
EPA/
OW/
OST/
HECD
Final
Draft
ppm
group,
and
from
53%
to
67%
in
the
150
ppm
exposure
group.
Decreased
food
consumption
(
reported
on
the
basis
of
grams/
kg
body
weight/
day)
was
observed
at
150
ppm
on
study
days
8
(
85%
of
controls)
and
10
(
89%
of
controls),
but
not
on
study
days
21
or
33.
A
14%
decrease
in
the
absolute
weight
of
the
right
cauda
epididymis
was
reported
for
the
50
ppm
males.
However,

the
absence
of
a
statistically
significant
effect
at
the
highest
concentration,
coupled
with
the
absence
of
effects
on
any
of
the
measured
sperm
parameters
or
on
epididymal
histopathology
suggests
that
this
result
is
not
biologically
significant.

Female
rats
were
divided
into
two
groups
to
separately
evaluate
toxicity
during
conception
and
early
gestation
(
designated
as
Group
A)
versus
toxicity
during
gestation
through
parturition
(
designated
as
Group
B).
Female
rats
for
both
groups
were
approximately
11
weeks
of
age
on
study
day
1.
Group
A
females
(
10
animals/
dose)
were
exposed
on
study
days
1
through
34,
cohabitated
with
treated
males
on
days
13
through
17,
and
examined
on
day
34
(
the
day
of
necropsy).
The
estimated
doses
resulting
from
exposure
to
0,
15,
50,
or
150
ppm
DBAN
were
0,

1.8,
5.1,
and
10.9
mg/
kg/
day
(
calculated
by
the
study
authors
from
body
weight
and
water
consumption
data).
Group
A
females
were
removed
from
cohabitation
upon
detection
of
vaginal
sperm
or
a
copulatory
plug,
or
after
five
days
in
the
absence
of
mating.
Clinical
signs
of
toxicity,

body
weight,
and
feed
and
water
consumption
were
determined
at
various
intervals.
At
study
termination
(
corresponding
to
gestation
day
16­
20),
gross
necropsy
was
performed
and
the
rats
were
evaluated
for
pregnancy
status,
number
and
position
of
live
and
dead
fetuses,
number
and
position
of
early
and
late
resorptions,
and
number
of
corpora
lutea.
Mating,
pregnancy,
and
fertility
indices,
total
number
of
implants,
and
pre­
and
post­
implantation
losses
were
calculated.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
29
EPA/
OW/
OST/
HECD
Final
Draft
No
treatment­
related
changes
were
observed
in
any
of
the
mating,
fertility,
pregnancy,
or
developmental
endpoints
that
were
examined.
All
of
the
maternal
parameters
evaluated
were
within
normal
limits,
except
for
a
7%
decrease
in
terminal
body
weight
(
non­
statistically
significant)
observed
at
150
ppm.
Feed
and
water
consumption
was
also
significantly
decreased
in
this
concentration
group.
Feed
consumption
was
statistically­
significantly
decreased
only
on
study
day
3
(
88%
of
controls)
in
the
150
ppm
exposure
group;
slight
decreases
in
food
consumption
on
other
study
days
were
not
statistically
significant.
Water
consumption
was
statistically­
significantly
decreased
in
the
150
ppm
exposure
group
on
all
four
study
days
for
which
the
data
were
summarized.
The
decrease
in
water
consumption
relative
to
controls
was
52%
on
day
3,
72%
on
day
5,
67%
on
day
21,
and
71%
on
day
33.
Slight
decreases
in
water
consumption
in
the
50
ppm
exposure
group
were
not
statistically
significant.

Group
B
females
(
13
animals/
dose)
were
cohabitated
with
males
beginning
on
study
day
1,

were
separated
as
soon
as
they
were
sperm­
positive,
had
a
copulatory
plug,
or
on
study
day
5,

and
then
were
exposed
on
gestation
day
6
through
postnatal
day
(
PND)
1.
The
estimated
doses
resulting
from
exposure
to
0,
15,
50,
or
150
ppm
DBAN
were
0,
1.9,
5.3,
and
10.8
mg/
kg/
day
(
calculated
by
the
study
authors
from
body
weight
and
water
consumption
data).
The
group
B
females
were
examined
on
PND
5,
and
pups
were
examined
on
PND
1,
3,
and
5.
Clinical
signs
of
toxicity,
body
weight,
and
feed
and
water
consumption
were
determined
at
various
intervals
during
the
study.
At
study
termination
(
PND
5),
females
were
examined
for
terminal
body
weight,
gross
necropsy
was
performed,
and
uterine
evaluation
was
conducted
with
the
number
of
implantation
sites
and
resorptions
recorded.
On
PND
1,
3,
and
5,
pup
weights,
number
of
live
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
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30
EPA/
OW/
OST/
HECD
Final
Draft
and
dead
pups,
and
number
of
male
and
female
pups
were
recorded.
The
anogenital
distance
of
pups
was
also
measured
on
PND
1.
All
parameters
were
within
normal
control
limits
except
for
decreased
water
consumption
in
the
50
and
150
ppm
exposure
groups
on
study
days
8,
14,
and
21,
but
not
on
study
day
6.
In
the
50
ppm
exposure
group
water
consumption
decreases
ranged
from
72%
to
83%
of
controls.
In
the
150
ppm
exposure
group
water
consumption
decreases
ranged
from
47%
to
60%.

This
study
suggests
that
DBAN
is
not
a
reproductive
or
developmental
toxicant
at
the
highest
dose
tested,
since
no
treatment­
related
effects
on
reproductive
or
developmental
parameters
were
observed
for
males
or
Group
A
or
Group
B
females.
However,
definitive
conclusions
regarding
the
potential
for
DBAN
to
induce
reproductive
or
developmental
effects
is
hampered
by
the
fact
that
this
was
a
screening
study
that
was
not
designed
to
evaluate
the
full
spectrum
of
endpoints
of
interest.
For
example,
males
were
not
exposed
during
all
stages
of
spermatogenesis,
and
the
pups
were
not
evaluated
for
possible
visceral
or
skeletal
malformations.

The
NOAEL
for
male
reproductive
effects
in
this
study
is
8.2
mg/
kg/
day.
The
NOAEL
for
female
reproductive
and
developmental
effects
is
10.8
mg/
kg/
day
(
the
lower
of
the
calculated
doses
for
Group
A
and
Group
B
females
exposed
to
the
highest
DBAN
concentration).
No
LOAEL
was
identified
for
male
or
female
reproductive
or
developmental
toxicity.
Slight
(
non­
statistically
significant)
body
weight
decreases
were
observed
in
high­
concentration
group
males
and
Group
A
females.
However,
these
results
were
not
considered
biologically
significant
because
of
their
limited
magnitude,
and
the
fact
that
they
might
be
secondary
to
decreased
water
consumption,

rather
than
due
to
a
toxicological
effect.
The
systemic
effects
evaluated
in
males
were
more
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
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31
EPA/
OW/
OST/
HECD
Final
Draft
extensive
than
in
females,
and
yield
a
critical
effect
level
with
the
greatest
degree
of
certainty.

Therefore,
the
NOAEL
of
8.2
mg/
kg/
day
for
males
is
selected
as
the
overall
study
NOAEL
for
systemic
toxicity.
No
LOAEL
was
identified.

Smith
and
colleagues
have
investigated
the
reproductive
and
developmental
effects
of
BCAN,
DBAN,
DCAN,
and
TCAN
through
a
series
of
studies,
first
testing
different
HANs
at
the
maximum
tolerated
dose,
and
then
conducting
dose­
response
studies
with
the
individual
compounds.
These
studies
were
followed
by
investigation
of
the
potentially
confounding
role
of
the
tricaprylin
solvent
vehicle
in
the
observed
developmental
toxicity.
NOAELs
and
LOAELs
are
reported
for
the
initial
studies,
conducted
in
tricaprylin
vehicle,
for
completeness.
However,
as
discussed
below
these
NOAELs
are
not
adequate
to
serve
as
the
basis
for
the
dose­
response
assessment,
since
tricaprylin
itself
contributes
to
the
developmental
toxicity
of
the
HANs,
and
the
effects
due
to
tricaprylin
cannot
be
separated
from
those
that
are
due
to
the
test
chemical.

In
the
initial
developmental
toxicity
screening
studies
(
Smith
et
al.,
1986;
Smith
et
al.,

1987)
sexually­
mature
Long­
Evans
rats
(
age
not
provided,
20
to
26
sperm
positive
rats
per
dose
group)
were
administered
gavage
doses
on
gestation
days
7
to
21
at
the
maximum
tolerated
dose
for
each
chemical
dissolved
in
tricaprylin:
55
mg/
kg/
day
for
BCAN,
DCAN,
and
TCAN;
50
mg/
kg/
day
for
DBAN.
Control
rats
received
tricaprylin
alone.
No
food
and
water
consumption
data
were
reported.
Duration
of
pregnancy,
litter
size,
sex
ratios,
and
litter
weights
were
determined
at
birth.
Dams
not
delivering
by
day
23
of
gestation
were
sacrificed,
and
the
status
of
their
pregnancies
was
determined.
Survival
and
weight
gain
of
offspring
to
PND
4
were
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
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32
EPA/
OW/
OST/
HECD
Final
Draft
measured.
At
PND
6,
litters
were
randomly
culled
to
a
consistent
number.
At
weaning,
the
number
of
pups
per
litter
were
further
randomly
reduced
to
a
consistent
number.
The
remaining
pups
were
monitored
for
growth
and
health
until
study
day
41­
42.
At
this
time,
all
surviving
pups
were
sacrificed
and
subjected
to
gross
necropsy,
with
livers,
kidneys,
spleens,
lungs,
and
gonads
removed
and
weighed.
The
following
parameters
were
evaluated:
maternal
toxicity
(
maternal
weight
gain
and
mortality),
reproductive
success
(
percent
pregnant,
percent
delivering
viable
litters),
and
growth
and
viability
of
pups
(
live
pups/
litter;
postnatal
survival;
mean
birth
weight;

survival,
weight
gain
and
development
through
necropsy;
gross
necropsy
and
organ
weight
results).
All
parameters
evaluated
were
within
normal
limits,
unless
described
below.

Treatment­
related
increases
in
maternal
deaths
were
observed
for
DBAN
(
15%
of
dams)

and
for
TCAN
in
one
study
group
(
20%
of
dams),
although
the
result
for
TCAN
was
not
replicated
in
a
second
study
group
receiving
the
same
dose.
The
study
authors
could
not
explain
the
disparate
results
in
duplicate
dose
groups
for
this
endpoint,
or
other
endpoints
noted
below.

Maternal
deaths
observed
in
the
BCAN
and
DCAN
groups
were
attributed
to
intubation
error.

No
maternal
deaths
were
reported
in
the
vehicle
controls.
All
four
HANs
caused
decreased
maternal
weight
gain
during
gestation,
although
the
decrease
was
not
statistically
significant
for
BCAN.
Since
the
maternal
weights
were
measured
prior
to
delivery,
decreased
unadjusted
maternal
weights
are
affected
by
litter
resorptions
and
decreased
pup
weights.
As
a
result,
the
observed
decreases
in
maternal
weight
gain
cannot
be
attributed
solely
to
maternal
toxicity.
Both
DCAN
and
TCAN
decreased
the
percentage
of
sperm­
positive
females
that
became
pregnant,
but
the
decrease
in
apparent
pregnancy
rate
was
observed
only
in
one
of
two
duplicate
dose
groups
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
33
EPA/
OW/
OST/
HECD
Final
Draft
for
each
compound.
According
to
the
study
authors,
the
apparent
decrease
in
pregnancy
rate
could
reflect
late
preimplantation
losses
or
very
early
resorptions.
This
conclusion
was
based
on
the
expected
stage
of
pregnancy
at
the
initiation
of
dosing
on
gestation
day
(
GD)
7
and
the
absence
of
detectable
ammonium­
sulfide
staining
decidua
(
which
would
indicate
implantation
sites).
DCAN
and
TCAN
(
in
only
one
of
two
duplicate
dose
groups
for
TCAN)
also
decreased
the
percentage
of
females
delivering
viable
litters
and
increased
the
percentage
of
litters
totally
resorbed
(
p

0.05).
Mean
birth
weights
were
reduced
for
all
four
compounds
(
p

0.05),
and
postnatal
survival
on
day
4
was
significantly
reduced
(
p

0.05)
in
pups
from
dams
exposed
to
DCAN
(
in
only
one
of
two
duplicate
dose
groups)
and
TCAN.
Cursory
visual
inspection
of
nonviable
pups
did
not
reveal
gross
terata,
although
the
authors
noted
that
the
current
study
protocol
was
not
expected
to
provide
a
sensitive
assessment
of
malformations
because
of
the
embryolethality
of
the
single
doses
that
were
tested.
No
pups
died
after
the
culling
of
litters
on
day
6
until
weaning,
but
after
weaning
some
pups
(
number
not
reported)
from
the
DCAN
replicate
died
because
they
were
too
small
to
reach
the
water
source.
Weight
gain
to
PND
4
was
significantly
decreased
for
male
and
female
pups
for
BCAN
and
DCAN
(
in
one
of
two
duplicate
dose
groups),
and
in
male
pups
for
DBAN.
Measurement
of
pup
weights
at
weaning
(
days
21­

22)
and
adolescence
(
days
41­
42)
revealed
significant
decreases
in
pup
body
weight
at
both
time
points
for
TCAN
and
at
adolescence
for
DCAN
(
in
one
of
two
duplicate
dose
groups).
Postnatal
weight
gain
from
weaning
until
puberty
and
sacrifice
was
reduced
in
both
male
and
female
pups
in
all
groups
administered
HANs,
but
these
effects
were
statistically
significant
only
in
males
and
females
receiving
TCAN
and
females
receiving
BCAN.
A
scattering
of
effects
in
the
relative
organ
to
body
weight
ratios
obtained
at
the
day
41­
42
sacrifice
was
reported
across
all
dose
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
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34
EPA/
OW/
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HECD
Final
Draft
groups.
However,
these
effects
are
judged
to
be
without
biological
significance
because
no
uniform
or
consistent
patterns
of
change
were
noted.

This
study
identifies
a
frank
effect
level
(
FEL)
of
50
mg/
kg/
day
for
DBAN
and
55
mg/
kg/
day
for
TCAN,
based
on
increased
maternal
deaths.
A
maternal
LOAEL
of
55
mg/
kg/
day
is
assigned
for
BCAN
and
DCAN
based
on
decreased
maternal
weight
gain.

However,
the
contribution
of
developmental
effects
(
e.
g.
decreased
litter
sizes,
fetal
weight)
to
the
observed
decrease
in
weight
gain
cannot
be
ruled
out,
and
therefore
it
is
not
known
if
the
observed
effect
was
due
to
systemic
toxicity,
severe
developmental
toxicity,
or
both.
The
LOAEL
for
developmental
effects
was
50
to
55
mg/
kg
for
all
four
compounds.
For
DCAN
and
TCAN,
these
LOAELs
were
based
on
severe
effects
(
increased
percent
of
early
resorptions
or
litters
totally
resorbed).
Maternal
and
developmental
NOAELs
were
not
identified
due
to
the
limited
number
of
test
doses
used
in
this
screening
assay.

In
a
follow­
up
study
evaluating
the
dose
response
for
TCAN,
Smith
et
al.
(
1988)

administered
TCAN
to
sperm­
positive
Long­
Evans
rats
aged
65­
80
days
(
19
to
21
per
dose
group
of
TCAN­
treated
animals,
30
rats
in
the
vehicle
control
group,
and
10
rats
in
water
controls)
by
gavage
in
tricaprylin
at
doses
of
0,
1.0,
7.5,
15,
35,
or
55
mg/
kg/
day
on
gestation
days
6
to
18.

Dams
were
sacrificed
on
day
21
of
gestation,
their
uterine
horns
were
examined
for
number
and
location
of
fetuses
or
resorption
sites,
and
the
fetuses
were
removed
and
examined.
Two­
thirds
of
each
litter
was
fixed
for
dissection
and
one­
third
stained
for
bone
and
cartilage
examination.

The
parameters
evaluated
in
this
study
were
maternal
toxicity
(
maternal
weight
gain
and
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
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35
EPA/
OW/
OST/
HECD
Final
Draft
mortality),
embryolethality
(
number
and
location
of
fetuses
or
resorption
sites,
number
of
viable
litters,
litter
sizes),
and
the
following
fetal
effects:
weights,
sex
ratios,
and
structural
abnormalities
including
external,
visceral,
and
skeletal
effects.
All
parameters
were
within
normal
limits,
unless
described
below
(
statistical
comparisons
are
made
to
the
tricaprylin
control
unless
noted
otherwise).
The
high
dose
was
lethal
in
4
out
of
19
dams,
and
maternal
weight
gain
was
decreased
beginning
at
15
mg/
kg/
day,
but
maternal
weight
gain
adjusted
for
fetal
weight
and
excluding
females
with
full­
litter
resorptions
was
significantly
decreased
only
at
55
mg/
kg/
day.

The
primary
developmental
effects
were
on
fetal
viability,
malformations,
and
fetal
body
weight.

There
was
a
dose­
related
increase
in
full­
litter
resorptions
(
compared
to
both
water
and
tricaprylin
control
groups)
at
7.5
mg/
kg/
day
and
higher,
affecting
2/
3
of
the
surviving
dams
at
the
high
dose.

Fetal
weight
was
significantly
decreased
only
at
35
mg/
kg/
day,
while
post­
implantation
loss
(
as
percent
resorptions
per
litter)
was
significantly
elevated
at
doses
of
15
mg/
kg/
day
and
higher.

Post­
implantation
losses
were
significantly
higher
and
male
fetal
body
weight
was
significantly
lower
(
p

0.05)
in
the
tricaprylin
controls
compared
to
the
water
controls.
The
authors
noted
that
the
value
obtained
for
post­
implantation
loss
for
water
controls
in
this
study
was
lower
than
the
historical
laboratory
control
levels
for
this
endpoint,
which
might
have
enhanced
the
observed
effect
of
the
tricaprylin
control.
While
no
malformations
were
observed
in
the
water
controls,

soft­
tissue
(
fetal
incidence
of
3.8%
and
litter
incidence
of
6/
30)
and
skeletal
(
fetal
incidence
of
13.3%
and
litter
incidence
of
7/
30)
malformations
were
observed
in
tricaprylin
controls.
The
percent
of
fetuses
affected
per
litter
with
soft­
tissue
malformations
was
dose­
dependent,
ranging
from
18%
at
7.5
mg/
kg/
day
to
35%
at
35
mg/
kg/
day.
This
value
decreased
to
22%
at
the
high
dose.
The
soft
tissue
malformation
frequency
at
the
low
dose
of
1.0
mg/
kg/
day
(
8.4%
of
fetuses
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
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36
EPA/
OW/
OST/
HECD
Final
Draft
affected
per
litter)
was
not
statistically
different
from
vehicle
controls
(
3.8%),
although
the
authors
expressed
concern
that
this
level
of
malformations
could
be
of
biological
significance.

However,
consideration
of
the
data
in
terms
of
the
more
appropriate
unit
(
U.
S.
EPA,
1991)
of
percent
of
litters
affected
indicated
no
effect
of
TCAN
compared
to
the
tricaprylin
control
(
4/
20
litters
affected
at
1
mg/
kg/
day
and
6/
30
litters
affected
in
tricaprylin
controls).
The
incidence
of
pups
with
cardiovascular
malformations
was
increased
at
15
mg/
kg/
day
and
urogenital
malformations
were
significantly
increased
at
both
15
mg/
kg/
day
and
35
mg/
kg/
day
(
p

0.05);
the
percentage
of
litters
with
cardiovascular
malformations
was
increased
beginning
at
7.5
mg/
kg/
day
(
8/
18
litters
affected
versus
6/
30
litters
in
tricaprylin
controls).
No
increase
in
the
incidence
of
external
or
skeletal
malformations
was
reported,
and
the
study
authors
do
not
appear
to
have
reported
on
the
incidences
of
external,
skeletal,
or
internal
variations
(
which
are
more
subtle
effects
than
malformations).

Since
the
apparent
increase
in
soft
tissue
malformations
at
1.0
mg/
kg/
day
was
not
statistically
significant
when
compared
to
tricaprylin
controls,
and
there
was
no
effect
on
the
percentage
of
affected
litters,
a
dose
of
1.0
mg/
kg/
day
of
TCAN
is
the
NOAEL
for
developmental
toxicity,
and
7.5
mg/
kg/
day
is
the
LOAEL,
although
this
conclusion
is
limited
by
the
absence
of
reported
data
on
the
incidence
of
variations.
The
maternal
NOAEL
for
this
study
is
35
mg/
kg/
day.
The
next
higher
dose
of
55
mg/
kg/
day
is
a
FEL,
based
on
increased
maternal
deaths.

Although
significant
decreases
in
overall
maternal
body
weight
gain
were
observed
beginning
at
15
mg/
kg/
day,
adjusted
body
weight
gain
was
decreased
only
at
55
mg/
kg/
day.
Adjusted
body
weight
gain
is
a
more
appropriate
indicator
of
maternal
toxicity
than
overall
body
weight
gain,
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
37
EPA/
OW/
OST/
HECD
Final
Draft
since
the
latter
would
be
affected
by
developmental
toxicity
as
well
as
maternal
toxicity.

Regardless
of
the
selection
of
the
critical
effect
levels,
the
results
of
this
study
are
not
appropriate
for
dose­
response
analysis
due
to
the
confounding
effects
of
the
tricaprylin
solvent
vehicle
as
evidenced
by
the
differences
between
water
and
tricaprylin
controls
in
this
study,
and
as
described
in
more
detail
in
the
study
by
Christ
et
al.
(
1996)
for
TCAN.

In
a
second
follow­
up
study
evaluating
the
dose­
response
for
DCAN,
Smith
et
al.
(
1989)

administered
DCAN
in
tricaprylin
to
pregnant
Long­
Evans
rats
aged
65­
80
days
(
22
to
24
rats
per
dose
group)
by
gavage
at
doses
of
0,
5,
15,
25,
or
45
mg/
kg/
day
on
gestation
days
6
to
18.

Tricaprylin
served
as
the
vehicle
control
and
distilled
water
served
as
an
additional
control
group.

Dams
were
sacrificed
on
day
20
of
gestation.
Their
livers,
spleens,
and
kidneys
were
removed
and
weighed,
and
their
uterine
horns
were
examined
for
number
and
location
of
fetuses
or
resorption
sites.
Fetuses
were
removed
and
examined.
Two­
thirds
of
each
litter
was
fixed
for
dissection
and
one­
third
stained
for
bone
and
cartilage
examination.
The
parameters
evaluated
in
this
study
were
maternal
toxicity
(
maternal
body
weight
gain,
liver,
spleen,
and
kidney
weight,
and
mortality),
embryolethality
(
number
and
location
of
fetuses
or
resorption
sites,
number
of
viable
litters,
litter
sizes),
and
the
following
fetal
effects:
weights,
crown­
rump
lengths,
sex
ratios,
and
structural
abnormalities
including
external,
visceral,
and
skeletal
effects.
All
parameters
were
within
normal
limits,
unless
described
below
(
statistical
comparisons
are
made
to
the
tricaprylin
control
unless
noted
otherwise).
Two
of
22
dams
died
in
the
high
dose
group
(
45
mg/
kg/
day).

No
maternal
deaths
were
observed
in
any
other
dose
group,
or
in
either
of
the
control
groups.

Maternal
body
weight
gain
and
adjusted
maternal
body
weight
gain
were
significantly
decreased
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
38
EPA/
OW/
OST/
HECD
Final
Draft
(
p

0.05)
at
45
mg/
kg/
day.
Surviving
dams
in
the
45
mg/
kg/
day
dose
group
showed
elevated
spleen
and
kidney
weights
(
p

0.05).
Liver
weight
was
significantly
increased
at
25
mg/
kg/
day,

but
not
at
the
high
dose.

DCAN
treatment
affected
a
number
of
developmental
toxicity
parameters.

Postimplantation
loss
was
significantly
increased
beginning
at
25
mg/
kg/
day.
At
45
mg/
kg/
day,
the
following
developmental
effects
were
observed:
increased
number
of
totally
resorbed
litters,

decreased
number
of
viable
litters,
decreased
fetal
weight
in
males
and
females,
and
decreased
crown­
rump
length
in
male
and
females.
The
incidence
of
malformations
increased
in
a
doserelated
manner.
Significant
increases
(
p

0.05)
in
total
soft
tissue,
cardiovascular,
urogenital
system,
and
skeletal
malformations
were
observed
in
fetuses
from
dams
exposed
to
45
mg/
kg/
day.

Increases
in
the
incidence
of
litters
affected
was
most
apparent
for
total
soft
tissue
malformations:

water
control
0/
19;
tricaprylin
control
4/
19;
5
mg/
kg/
day
5/
22;
15
mg/
kg/
day
5/
16;
25
mg/
kg/
day
9/
16;
45
mg/
kg/
day
7/
7.
The
authors
did
not
present
a
statistical
analysis
of
the
malformation
data
presented
as
the
incidence
of
litters
affected.
Analysis
of
these
data
for
preparation
of
this
Criteria
Document
(
using
Fischer's
Exact
Test)
revealed
that
the
litter
incidence
of
malformations
is
significantly
greater
(
P<
0.05)
in
the
25
mg/
kg/
day
and
45
mg/
kg/
day
dose
groups
than
in
tricaprylin
controls.
The
water
controls
were
not
significantly
different
than
the
tricaprylin
controls
(
P
=
0.0525),
based
on
the
litter
incidence,
although
the
study
authors
reported
a
significant
difference
between
water
and
tricaprylin
controls
on
the
basis
of
fetuses
affected
per
litter.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
39
EPA/
OW/
OST/
HECD
Final
Draft
The
NOAEL
for
maternal
toxicity
for
this
study
is
15
mg/
kg/
day
and
the
LOAEL
is
25
mg/
kg/
day
based
on
increased
liver
weight.
No
additional
measures
of
liver
toxicity
were
included
in
this
study,
but
the
selection
of
increased
liver
weight
as
an
adverse
effect
is
supported
by
ability
of
DCAN
to
induce
liver
toxicity
at
similar
doses
as
reported
in
the
subacute
and
subchronic
studies
by
Hayes
et
al.
(
1986).
The
NOAEL
for
developmental
toxicity
is
15
mg/
kg/
day
and
the
LOAEL
is
25
mg/
kg/
day,
based
on
increased
post­
implantation
loss
and
malformations.
No
differences
in
fetal
viability
or
size
were
noted
between
the
water
and
tricaprylin
control
groups.
However,
the
incidence
of
total
soft
tissue
malformations
was
significantly
lower
in
water
(
0/
19
litters
affected)
versus
tricaprylin
controls
(
4/
19
litters
affected),
suggesting
that
tricaprylin
induces
developmental
toxicity,
and
that
this
effect
is
further
potentiated
following
combined
treatment
with
tricaprylin
and
DCAN
(
as
high
as
9/
16
litters
affected).
Such
a
finding
is
of
considerable
importance
because,
as
mentioned
in
the
paragraph
above
introducing
the
Smith
studies,
it
raises
the
possibility
that
the
fetal
malformations
attributed
to
the
HANs
in
this
study
may
result
from
the
potentiation
by
the
tricaprylin
vehicle.
Tricaprylin
effects
and
their
implications
for
assessing
HAN
toxicity
are
discussed
further
in
the
studies
below.

The
developmental
toxicity
of
BCAN
was
evaluated
in
120
to
150­
day
old
Long­
Evans
rats
(
Christ
et
al.,
1995)
in
a
third
study
to
evaluate
the
dose­
response
for
HANs.
Pregnant
rats
were
administered
BCAN
by
gavage
in
tricaprylin
on
gestation
days
6
to
18
at
doses
of
0,
5,
25,

45,
or
65
mg/
kg/
day.
Dams
treated
with
tricaprylin
only
served
as
the
vehicle
control
and
dams
treated
with
distilled
water
served
as
an
additional
control.
The
number
of
rats
per
treatment
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
40
EPA/
OW/
OST/
HECD
Final
Draft
group
is
unclear,
because
while
the
methods
section
indicates
that
between
17
and
23
animals
were
assigned
per
dose
group,
results
for
reproductive
performance
presented
in
the
paper
suggest
group
sizes
of
20
to
28
dams.
Dams
were
sacrificed
on
day
20
of
gestation.
Their
livers,

spleens,
and
kidneys
were
removed
and
weighed,
and
their
uterine
horns
were
examined
for
number
and
location
of
fetuses
or
resorption
sites.
Fetuses
were
removed
and
examined.

Twothirds
of
each
litter
was
fixed
for
dissection
and
one­
third
stained
for
bone
and
cartilage
examination.
The
parameters
evaluated
in
this
study
were
maternal
toxicity
(
maternal
body
weight
gain,
liver,
spleen,
and
kidney
weight,
and
mortality),
embryolethality
(
number
and
location
of
fetuses
or
resorption
sites,
number
of
viable
litters,
litter
sizes),
and
the
following
fetal
effects:
weights,
crown­
rump
lengths,
sex
ratios,
and
structural
abnormalities,
including
external,

visceral,
and
skeletal
effects.
All
parameters
were
within
normal
limits,
unless
described
below
(
statistical
comparisons
are
made
to
the
tricaprylin
control
unless
noted
otherwise).

Treatment
with
BCAN
in
tricaprylin
resulted
in
both
maternal
and
developmental
toxicity.

Mortality
was
statistically
significantly
increased
in
the
high­
dose
dams
(
3
of
26
treated
dams)

compared
with
the
water
and
tricaprylin
control
groups
in
which
no
treatment­
related
deaths
were
observed.
Dams
in
both
the
45
and
65
mg/
kg/
day
dose
groups
had
significantly
decreased
percentage
of
body
weight
gain
compared
with
the
tricaprylin
control
group,
and
dams
in
the
65
mg/
kg/
day
group
had
decreased
body
weight
gain
after
adjusting
for
gravid
uterine
weight.

Kidney
weights
were
significantly
increased
in
dams
at
doses

25
mg/
kg/
day
compared
with
tricaprylin
controls.
Liver
and
spleen
weights
were
significantly
increased
only
in
the
high­
dose
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
41
EPA/
OW/
OST/
HECD
Final
Draft
dams.
There
was
no
difference
in
any
of
these
maternal
toxicity
parameters
between
tricaprylin
and
water
controls.

In
terms
of
reproductive
and
developmental
endpoints,
an
increase
in
the
number
of
litters
totally
resorbed
and
a
decrease
in
the
number
of
viable
litters
compared
with
tricaprylin
controls
were
observed
beginning
at
45
mg/
kg/
day.
The
percent
post­
implantation
loss
and
the
percent
of
resorbed
fetuses
per
litter
increased
at
doses
of

45
mg/
kg/
day
compared
with
the
tricaprylin
control
group.
Although
the
increase
for
post­
implantation
loss
was
not
statistically
significant
in
the
high­
dose
group,
this
parameter
was
statistically
significant
in
the
second­
highest
dose
group,

and
also
substantially
elevated,
although
not
statistically,
in
the
third­
highest
dose
group.
Fetal
crown­
rump
length
was
significantly
decreased
in
all
the
treated
groups,
and
fetal
weights
were
decreased
beginning
at
25
mg/
kg/
day.
Statistical
analysis
of
the
malformation
data
was
done
in
terms
of
the
percent
of
fetuses
affected
per
litter.
However,
examination
of
the
data
on
the
basis
of
the
percent
of
litters
affected
appears
to
lead
to
similar
conclusions
with
regard
to
selection
of
effect
levels.
A
significant
increase
in
cardiovascular
malformations
compared
with
tricaprylin
controls
was
observed
in
all
the
dose
groups.
Urogenital
malformations
were
significantly
increased
only
at
45
mg/
kg/
day,
although
the
percentage
of
litters
affected
appeared
to
be
significantly
greater
at
both
25
and
45
mg/
kg/
day.
Total
soft
tissue
malformations
were
increased
beginning
at
25
mg/
kg/
day
and
skeletal
malformations
were
significantly
increased
beginning
at
45
mg/
kg/
day.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
42
EPA/
OW/
OST/
HECD
Final
Draft
In
addition
to
the
BCAN­
treated
groups,
tricaprylin
vehicle
alone
had
significant
effects
on
embryotoxicity.
Developmental
endpoints
affected
by
tricaprylin
are
noted
as
mean
±
standard
error
tricaprylin
versus
water
controls.
Animals
in
the
tricaprylin
control
group
had
significantly
increased
percent
post­
implantation
loss
(
15.5
±
19.0
versus
6.7
±
9.8);
decreased
fetal
body
weight
(
grams)
in
males
(
3.19
±
0.3
versus
3.61
±
0.2);
decreased
fetal
body
weight
(
grams)
in
females
(
2.90
±
0.3
versus
3.41
±
0.2);
decreased
crown­
rump
length
(
cm)
in
males
(
3.4
±
0.2
versus
3.6
±
0.2);
and
decreased
crown­
rump
length
(
cm)
in
females
(
3.4
±
0.2
versus
3.5
±
0.1).

An
increased
incidence
of
urogenital
malformations
(
0/
19
litters
versus
4/
23
litters)
was
also
induced
by
tricaprylin
as
compared
to
water
controls.

The
NOAEL
for
maternal
toxicity
in
this
study
is
45
mg/
kg/
day
and
the
high
dose
of
65
mg/
kg/
day
is
a
FEL
based
on
increased
maternal
deaths,
and
accompanied
by
decreased
adjusted
maternal
weight
gain,
and
organ
weight
changes.
The
increase
in
kidney
weight
at
the
lower
doses
was
not
used
as
the
basis
for
assigning
maternal
toxicity
effect
levels
because
of
the
absence
of
data
to
determine
whether
this
effect
was
adverse.
Developmental
effects,
including
decreased
crown­
rump
length
and
increased
cardiovascular
malformations
were
observed
at
the
the
low
dose
of
5
mg/
kg/
day.
Therefore,
5
mg/
kg/
day
is
a
developmental
LOAEL
for
this
study,
and
no
NOAEL
is
identified.
However,
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),
use
of
this
study
for
dose­
response
assessment
is
not
appropriate,
because
it
may
not
accurately
reflect
the
toxicity
of
BCAN
in
drinking
water
in
the
absence
of
the
tricaprylin
vehicle.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
43
EPA/
OW/
OST/
HECD
Final
Draft
Based
on
the
observed
increased
embryotoxicity
in
tricaprylin
versus
water­
treated
controls
in
earlier
studies,
Christ
et
al.
(
1996)
investigated
the
effect
of
solvent
vehicle
on
the
developmental
toxicity
of
TCAN
(
Table
V­
10).
Groups
of
approximately
20
sperm­
positive
Long­
Evans
rats
aged
65­
80
days
were
treated
with
15,
35,
55,
or
75
mg/
kg/
day
TCAN
in
corn
oil,
or
15
mg/
kg/
day
TCAN
in
tricaprylin.
In
addition,
water,
corn
oil,
and
tricaprylin
were
used
as
controls.
Treatments
were
administered
by
oral
gavage
on
gestation
days
6
to
18
for
all
of
the
treatment
groups.
Dams
were
sacrificed
on
day
20
of
gestation.
Their
livers,
spleens,
and
kidneys
were
removed
and
weighed,
and
their
uterine
horns
were
examined
for
number
and
location
of
fetuses
or
resorption
sites.
Fetuses
were
removed
and
examined.
Two­
thirds
of
each
litter
was
fixed
for
dissection
and
one­
third
stained
for
bone
and
cartilage
examination.
The
parameters
evaluated
in
this
study
were
maternal
toxicity
(
maternal
body
weight
gain,
liver,
spleen,
and
kidney
weight,
and
mortality),
embryolethality
(
number
and
location
of
fetuses
or
resorption
sites,

number
of
viable
litters,
litter
sizes),
and
the
following
fetal
effects:
weights,
crown­
rump
lengths,

sex
ratios,
and
structural
abnormalities,
including
external,
visceral,
and
skeletal
effects.
All
parameters
were
within
normal
limits,
unless
described
below.
Of
the
20
dams
treated
with
75
mg/
kg/
day
TCAN
in
corn
oil,
five
dams
died,
five
were
nonpregnant,
and
nine
dams
resorbed
their
entire
litter,
so
that
only
one
viable
litter
was
produced.
For
this
reason,
other
data
for
the
75
mg/
kg/
day
group
were
not
reported.
No
maternal
deaths
were
reported
in
any
of
the
other
groups.
Maternal
weight
gain
was
significantly
decreased
beginning
at
15
mg/
kg/
day,
and
maternal
weight
gain,
after
adjusting
for
gravid
uterine
weight,
was
significantly
decreased
at
35
mg/
kg/
day
and
higher
doses.
Relative
maternal
liver
weight
was
increased
at

35
mg/
kg/
day,
and
liver,
spleen,
and
kidney
weights
were
significantly
increased
at
55
mg/
kg/
day
as
compared
to
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
44
EPA/
OW/
OST/
HECD
Final
Draft
corn
oil
vehicle
controls.
Similarly,
increased
relative
liver
weights
were
observed
in
the
animals
given
15
mg/
kg/
day
TCAN
in
tricaprylin
as
compared
to
15
mg/
kg/
day
TCAN
administered
in
corn
oil.

The
percent
post­
implantation
loss
was
significantly
increased,
and
the
number
of
live
fetuses
per
litter,
fetal
body
weight,
and
crown­
rump
length
were
all
significantly
decreased
in
the
group
administered
55
mg/
kg/
day
TCAN
in
corn­
oil
vehicle.
Also
in
this
dose
group,
the
mean
percentage
of
fetuses
per
litter
with
external
malformations,
skeletal
malformations,
and
softtissue
malformations
was
significantly
increased.
The
incidence
of
cardiovascular
and
urogenital
malformations
was
not
increased
at
any
dose,
but
other
soft
tissue
malformations
(
i.
e.
not
classified
as
either
cardiovascular
or
urogenital)
were
significantly
increased
in
the
55
mg/
kg/
day
dose
group.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
Final
Draft
V
­
45
Table
V­
10.
Reproductive
and
Developmental
Toxicity
of
TCANa.

TCAN
(
mg/
kg/
d
)
TCAN
(
mg/
kg/
d)

Water
Corn
oil
15
35
55
75
Tricaprylin
15
No.
Sperm­
positive
treated
females
20
20
17
21
21
20
20
21
Nonpregnant
females
2
3
0
0
3
5
1
2
Deaths
0
0
0
0
0
5b
0
0
Litters
totally
resorbed
0
0
0
0
1
9b
0
3
Viable
littersc
18
17
17
21
17
1
19
16
Percent
maternal
weight
gaind
45.5
±
9.0
48.9
±
8.1
44.2
±
7.8b
36.9
±
8.8b
23.7
±
7.7b
nde
41.6
±
10.0
41.7
±
10.5
Adjusted
maternal
weight
gaind
18.6
±
5.5
20.4
±
4.8
17.6
±
5.6
11.1
±
5.7b
6.8
±
5.4b
nde
18.1
±
5.8
17.3
±
6.7
Total
implants
per
litter
13.6
±
1.6
14.2
±
1.3
13.8
±
1.3
13.4
±
2.4
13.8
±
1.7
12.3
±
4.3
12.7
±
2.5f
13.4
±
3.0
Percent
preimplantation
lossg
3.3
±
5.6
4.6
±
7.5
3.5
±
6.6
5.0
±
13.8
3.8
±
8.3
17.7
±
28.1
14.0
±
13.9h
11.6
±
20.0
Percent
post­
implantation
lossi
12.1
±
12.4
7.0
±
7.6
6.3
±
7.0
8.4
±
11.1
29.7
±
25.2b
98.8
±
4.0b
15.9
±
17.8
25.4
±
34.8
Live
fetuses
per
litter
11.9
±
2.1
13.2
±
1.6
12.9
±
1.4
12.3
±
2.8
10.4
±
2.6b
nd
10.7
±
3.2f
12.4
±
1.8
Fetal
body
weight
(
g)
(
male)
3.58
±
0.25
3.42
±
0.17
3.52
±
0.20
3.38
±
0.30
2.54
±
0.46b
nd
3.22
±
0.28f
2.93
±
0.37b,
j
Fetal
body
weight
(
g)
(
female)
3.41
±
0.26
3.28
±
0.19
3.26
±
0.22
3.25
±
0.30
2.39
±
0.41b
nd
3.01
±
0.25h
2.76
±
0.38b,
j
Crown­
rump
length
(
cm)

(
male)
3.58
±
0.08
3.57
±
0.11
3.60
±
0.13
3.54
±
0.14
3.28
±
0.23b
nd
3.45
±
0.13h
3.46
±
0.16j
Drinking
Water
Criteria
Document
for
Haloacetonitriles
Table
V­
10.
Reproductive
and
Developmental
Toxicity
of
TCANa.

TCAN
(
mg/
kg/
d
)
TCAN
(
mg/
kg/
d)

Water
Corn
oil
15
35
55
75
Tricaprylin
15
EPA/
OW/
OST/
HECD
Final
Draft
V
­
46
Crown­
rump
length
(
cm)

(
female)
3.51
±
0.09
3.52
±
0.09
3.51
±
0.17
3.45
±
0.13
3.16
±
0.19b
nd
3.36
±
0.13h
3.36
±
0.18j
Fetal
malformations
External
0
0
0
1.4
±
6.2
(
1/
21)
k
5.3
±
8.8b
(
6/
17)
nd
0
0.5
±
1.9
(
1/
16)

Total
soft
tissue
(
visceral)
0
1.9
±
4.3
(
3/
17)
0.6
±
2.4
(
1/
17)
0
15.0
±
17.1b
(
3/
17)
nd
2.4
±
6.3
(
3/
19)
15.0
±
21.2b,
j
(
9/
16)

Cardiovascular
0
0.6
±
2.4
(
1/
17)
0.6
±
2.4
(
1/
17)
0
4.5
±
11.2
(
3/
17)
nd
1.3
±
5.7
(
1/
19)
12.0
±
17.3b,
j
(
7/
16)

Urogenital
0
1.3
±
3.8
(
2/
17)
0
0
0
nd
2.4
±
6.3
(
3/
19)
3.0
±
7.1
(
4/
16)

Other
soft
tissues
0
0
0
0
10.4
±
16.3b
(
3/
17)
nd
0
0
Skeletal
0
0
0
1.6
±
7.3
(
1/
21)
7.1
±
14.4b
(
2/
17)
nd
0
0
a.
Adapted
from
Christ
et.
al.
(
1996)

b.
Significantly
different
from
vehicle
control,
p

0.05.

c.
Viable
litters
were
those
containing
at
least
one
live
pup.

d.
Weight
gain
analysis
included
only
females
with
viable
litters
(
mean
±
SD
reported);
adjusted
maternal
weight
gain
controls
for
the
effect
of
gravid
uterine
weight.

e.
Values
for
75
mg/
kg/
day
are
not
reported
since
there
was
only
one
dam
with
a
viable
litter.

f.
Significantly
different
from
corn
oil
control,
p

0.05.
g.
Percent
preimplantation
loss
=
(
number
of
copora
lutea
­
number
of
implants)
/
number
of
corpora
lutea
x
100.

h.
Significantly
different
from
water
or
corn
oil
control,
p

0.05.

i.
Percent
post­
implantation
loss
=
(
number
of
implants
­
number
of
live
fetuses)
/
number
of
implants
x
100.

j.
Significantly
different
from
TCAN
15/
mg/
kg/
day
in
corn
oil.

k.
(
number
litters
examined
/
number
of
litters
affected)
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
47
EPA/
OW/
OST/
HECD
Final
Draft
No
developmental
effects
were
observed
at
the
doses
below
55
mg/
kg/
day
of
TCAN
in
corn
oil.
By
contrast,
fetal
body
weight
was
significantly
decreased
in
the
group
dosed
with
15
mg/
kg/
day
TCAN
in
the
tricaprylin
vehicle
and
the
percentage
of
fetuses
per
litter
with
total
soft
tissue
and
cardiovascular
malformations
also
was
significantly
increased
in
this
group
(
both
comparisons
relative
to
both
the
the
tricaprylin
control
and
the
same
dose
of
TCAN
in
corn
oil).

The
authors
noted
a
shift
in
the
spectrum
of
soft
tissue
malformations
from
cardiovascular
(
communication
or
vascular
defects)
and
urogenital
effects
in
the
groups
treated
with
TCAN
in
tricaprylin
to
external
cranio­
facial
malformations
and
positional
cardiovascular
malformations
(
e.
g.,
levocardia)
in
the
groups
treated
with
TCAN
in
corn
oil.
Comparing
the
two
vehicles,

TCAN
administered
at
15
mg/
kg/
day
in
tricaprylin
produced
effects,
including
increased
liver
and
kidney
weights,
decreased
fetal
weight,
decreased
crown­
rump
length,
and
increased
percent
of
fetuses
with
soft­
tissue
malformations
that
were
not
observed
when
TCAN
was
administered
at
15
mg/
kg/
day
in
corn
oil.
When
the
water,
corn
oil,
and
tricaprylin
control
groups
were
compared,
no
differences
were
observed
between
the
water
and
corn
oil
groups.
However,
the
tricaprylin
group
had
the
following
statistically
significant
changes
compared
to
water
or
corn
oil
groups:
increased
maternal
kidney
weight,
decreased
total
implants
per
litter,
increased
preimplantation
loss,
decreased
live
fetuses/
litter,
and
decreased
fetal
weight
and
crown­
rump
length.

The
comparison
of
solvent
vehicle
effects
in
this
study
clearly
shows
that
when
TCAN
is
administered
in
tricaprylin,
maternal
and
developmental
toxicity
are
observed
at
lower
doses
and
the
spectrum
of
effects
is
changed
as
compared
to
TCAN
administered
in
corn
oil.
Since
corn
oil
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
48
EPA/
OW/
OST/
HECD
Final
Draft
and
water
control
responses
were
not
different
in
this
study,
but
tricaprylin
controls
showed
increased
toxicity
for
several
endpoints,
the
data
for
TCAN
dissolved
in
corn
oil
are
judged
to
be
most
relevant
for
dose­
response
assessment,
and
the
data
for
TCAN
in
tricaprylin
are
judged
as
inadequate
for
use
in
dose­
response
assessment.
Based
on
this
rationale,
the
NOAEL
for
maternal
toxicity
in
this
study
is
15
mg/
kg/
day,
and
the
LOAEL
is
35
mg/
kg/
day
for
decreased
maternal
body
weight
after
adjusting
for
gravid
uterine
weight.
No
developmental
effects
were
observed
below
55
mg/
kg/
day
TCAN
in
corn
oil.
Therefore,
for
developmental
effects,
the
NOAEL
of
TCAN
in
corn
oil
is
35
mg/
kg/
day
and
the
LOAEL
is
55
mg/
kg/
day.
The
data
sets
for
maternal
and
developmental
endpoints
were
further
analyzed
to
determine
benchmark
doses
(
BMDs)
according
to
draft
EPA
Guidance
(
U.
S.
EPA,
2000c)
to
identify
alternative
critical
effect
levels.
The
results
of
the
modeling
are
described
in
detail
in
Appendix
A.
A
BMDL
of
17
mg/
kg/
day
for
decreased
adjusted
maternal
body
weight
gain
was
selected
as
the
most
appropriate
modeling
result
to
serve
as
the
basis
for
the
quantitative
dose­
response
assessment.

Based
on
the
effects
of
tricaprylin
alone
and
its
ability
to
potentiate
the
toxicity
of
TCAN,

the
use
of
the
data
from
the
developmental
toxicity
studies
using
this
vehicle
is
not
appropriate
for
risk
assessment.
The
mechanism
responsible
for
the
greater
sensitivity
and
different
pattern
of
malformations
produced
by
TCAN
when
it
is
administered
in
tricaprylin
instead
of
water
or
corn
oil
is
not
understood.
However,
an
earlier
abstract
by
this
same
laboratory
(
Gordon
et
al.,
1991)

suggested
that
multiple
exposures
to
TCAN
in
tricaprylin
results
in
different
distribution
of
TCAN
(
or
the
TCAN/
tricaprylin
combination)
to
maternal
tissues
and
embryos
than
occurs
when
TCAN
is
administered
in
corn
oil.
However,
as
described
in
more
detail
in
Chapter
III
(
Toxicokinetics)
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
49
EPA/
OW/
OST/
HECD
Final
Draft
the
data
are
not
sufficient
to
determine
whether
the
observed
differences
in
response
when
tricaprylin
was
used
as
the
solvent
vehicle
were
due
to
a
toxicokinetic
or
a
toxicodynamic
interaction.

D.
Mutagenicity
and
Genotoxicity
The
genotoxicity
of
HANs
has
been
tested
in
a
variety
of
diverse
assays.
In
this
section
we
present
the
study
results
for
the
HANs
by
study
type.
The
results
of
mutagenicity
assays
are
described
first,
followed
by
measures
of
chromosome
effects,
and
then
assays
of
DNA
damage.

Table
V­
12
provided
at
the
end
of
this
section
summarizes
the
overall
genotoxicity
study
results
on
a
chemical­
by­
chemical
basis.

The
mutagenicity
of
HANs
has
been
studied
by
several
investigators
using
standard
protocols
or
variations
of
the
Salmonella
typhimurium/
mammalian
microsome
mutagenesis
assay.

In
a
report
summarizing
results
of
the
USEPA
Gene­
Tox
program,
Kier
et
al.
(
1986)
presented
the
results
of
testing
DCAN
in
strains
TA1535,
TA1537,
TA1538,
TA100,
and
TA98.
These
varying
tester
strains
are
used
to
identify
a
range
of
mutagenic
target
sites,
where
strains
TA100
and
TA1535
detect
transitions
and
transversions,
while
strains
TA98
and
TA1538
detect
frameshifts
and
small
deletions/
insertions.
(
The
study
results
were
initially
reported
by
Simmon
et
al.,
1977;
and
Simmon
and
Kauhanen,
1978).
A
positive
result
was
defined
as
the
generation
of
greater
revertant
counts
than
controls
at
two
doses
and
at
one
dose
the
number
of
the
revertants
had
to
be
3­
times
greater
than
control
for
strains
TA1535,
TA1537,
or
TA1538,
or
twice
the
control
value
for
TA100
or
TA98.
Positive
mutagenic
activity
with
or
without
addition
of
S9
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
50
EPA/
OW/
OST/
HECD
Final
Draft
activation
was
reported
in
strains
TA100
and
TA1535,
while
findings
were
negative
regardless
of
S9
activation
in
strains
TA98,
TA1537,
and
TA1538.
Overall,
DCAN
was
positive
in
this
assay.

Bull
et
al.
(
1985)
studied
the
mutagenic
effects
of
BCAN,
DBAN,
DCAN,
and
TCAN
in
the
Salmonella/
microsome
assay
using
tester
strains
TA98,
TA100,
TA1535,
TA1537,
and
TA1538.
Cells
were
dosed
with
up
to
5.44
µ
mol/
plate
BCAN,
up
to
0.58
µ
mol/
plate
DBAN,
up
to
12.4
µ
mol/
plate
of
DCAN,
and
up
to
11.7
µ
mol/
plate
TCAN.
Wide
ranges
of
doses
were
used
and
the
highest
concentration
equaled
or
exceeded
the
LC
50
for
each
compound.
In
this
assay,

BCAN
was
positive
in
strain
TA100
(+
S9)
and
in
TA1535
(+
S9
or
­
S9),
while
DCAN
was
positive
in
TA98,
TA100,
and
TA1535
regardless
of
S9
activation.
The
positive
results
for
BCAN
and
DCAN
in
strain
TA1535
led
the
authors
to
conclude
that
both
of
these
compounds
can
induce
base­
pair
substitutions.
None
of
the
HANs
increased
the
number
of
revertants
in
strains
TA1537
or
TA1538.
In
addition,
DBAN
and
TCAN
failed
to
produce
dose­
related
increases
in
the
frequency
of
histidine
revertants
in
any
strain.

The
Ames­
fluctuation
test
was
performed
by
exposing
S.
typhimurium
strain
TA100
to
the
HANs
in
liquid
culture
with
and
without
the
addition
of
S9
(
Le
Curieux
et
al.,
1995).
The
range
of
doses
tested,
in
µ
g/
mL,
were
as
follows:
BCAN,
0.03­
10
(­
S9),
0.3­
100
(+
S9);
DBAN,
0.03­

10
(­
S9),
0.1­
30
(+
S9);
DCAN,
0.3­
300
(­
S9),
0.3­
1000
(+
S9);
TCAN,
0.1­
1000
(+/­
S9).
In
all
cases
the
tested
dose
range
included
cytotoxic
concentrations
(
defined
as
sufficient
to
reduce
the
number
of
positive
wells
compared
to
controls).
All
of
the
compounds
except
DBAN
generated
positive
results
(
i.
e.
generated
a
statistically
significant
increase
in
the
number
of
positive
wells
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
51
EPA/
OW/
OST/
HECD
Final
Draft
compared
to
controls).
The
effective
doses
ranged
widely
and
no
clear
pattern
of
dependence
on
S9
activation
was
apparent.
BCAN
yielded
a
positive
result
at
a
concentration
as
low
as
0.6
µ
g/
mL
(­
S9),
DCAN
yielded
a
positive
result
at
10
µ
g/
mL
(­
S9)
and
300
µ
g/
mL
(+
S9),
and
TCAN
at
30
µ
g/
mL
(­
S9).

Gee
et
al.
(
1998)
conducted
a
validation
analysis
for
the
Salmonella/
microsome
assay
by
comparing
the
effects
of
a
group
of
substances
in
base­
specific
tester
strains
to
mutagenic
activity
in
traditional
strains.
Mutagenic
activity
of
TCAN
in
the
base­
specific
strains
TA7001,
TA7002,

TA7003,
TA7004,
TA7005,
and
TA7006
and
a
mix
of
these
six
strains
was
contrasted
to
the
mutagenic
activity
of
TCAN
in
the
traditional
frameshift
strains
TA98
and
TA1537.
The
assay
was
done
using
a
liquid
fluctuation
exposure
protocol
in
the
presence
and
absence
of
S9
activation.
For
each
combination
of
tester
strains,
four
doses
of
TCAN
ranging
from
50
to
1000
µ
g/
mL
in
dimethyl
sulfoxide
were
used,
plus
solvent
control
and
positive
control
groups.
The
test
agent
was
considered
mutagenic
if
any
of
the
test
doses
were
found
to
generate
revertants
at
statistically
significant
levels
compared
to
the
control.
The
only
positive
result
reported
was
for
the
base­
specific
strain
mix
with
S9
activation.
This
result
must
be
interpreted
with
caution,

however,
because
the
authors
did
not
explain
why
a
positive
result
was
identified
in
the
basespecific
strain
mix,
when
none
of
the
base­
specific
strains
was
positive
when
tested
individually.

In
addition,
inspection
of
the
raw
data
(
available
from
an
online
site
provided
in
the
paper)
did
not
reveal
any
clear
increase
in
the
frequency
of
revertants
in
the
mix
versus
the
individual
test
strains.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
52
EPA/
OW/
OST/
HECD
Final
Draft
Taken
together,
these
mutagenicity
assays
indicate
that
BCAN
and
DCAN
are
mutagenic
in
S.
typhimurium
(
Bull
et
al.,
1985;
Kier
et
al.,
1986;
Le
Curieux
et
al.,
1995).
TCAN
has
produced
mixed
results,
with
negative
results
reported
in
the
standard
assay
protocol
(
Bull
et
al.,

1985;
Le
Curieux
et
al.,
1995;
Gee
et
al.,
1998).
DBAN
has
yielded
negative
results
in
all
of
the
assays
reported
(
Bull
et
al.,
1985;
Le
Curieux
et
al.,
1995).
No
studies
testing
any
of
these
compounds
in
mammalian
gene
mutation
assays
were
located.

Several
investigators
have
tested
the
ability
of
HANs
to
induce
chromosome
damage,
with
mixed
results.
Bull
et
al.
(
1985)
studied
the
ability
of
HANs
to
produce
chromosomal
damage
or
loss
by
examining
micronuclei
production
in
polychromatic
erythrocytes
in
an
in
vivo
assay
in
CD­

1
mice
(
5
animals/
sex/
group).
Animals
were
dosed
by
gavage
with
BCAN,
DBAN,
DCAN,
or
TCAN
dissolved
in
10%
Emulphor
at
0,
12.5,
25,
or
50
mg/
kg/
day
for
five
consecutive
days
and
sacrificed
6
hours
after
the
last
dose.
The
highest
dose
was
selected
to
generate
a
cumulative
dose
(
5
doses
x
50
mg/
kg/
day
=
250
mg/
kg/
day)
approximating
the
oral
LD
50.
No
significant
increases
in
micronuclei
frequency
were
observed
for
any
of
the
HANs
tested.
It
is
unclear,

however,
whether
sufficiently
high
doses
were
tested
in
this
study.
The
study
authors
did
not
present
the
supporting
data
and
did
not
report
whether
cytotoxicity
of
the
target
tissue
occurred
(
as
evidenced
by
a
change
in
the
ratio
of
polychromatic
to
normochromatic
erythrocytes).
In
addition,
the
shallow
duration­
response
curve
seen
for
the
14­
day
versus
90­
day
toxicity
(
Hayes
et
al.,
1986)
suggests
that
the
cumulative
dose
is
not
an
appropriate
surrogate
for
a
single
dose
at
the
LD
50.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
53
EPA/
OW/
OST/
HECD
Final
Draft
In
contrast
to
this
result,
other
investigators
have
reported
positive
evidence
for
chromosome
damage
using
less
standard
assay
systems.
Le
Curieux
et
al.
(
1995)
treated
Pleurodeles
waltl
larvae
(
newt)
with
HANs
dissolved
in
the
container
water.
Following
12
days
of
treatment,
blood
was
collected
and
micronucleated
erythrocytes
were
counted
from
a
population
of
1,000
erythrocytes.
The
range
of
concentrations
tested,
in
µ
g/
mL,
were
as
follows:

BCAN,
0.0312­
0.125;
DBAN,
0.12­
0.5;
DCAN,
0.25­
1;
TCAN,
0.025­
0.1.
All
the
test
compounds
generated
a
positive
result
in
this
assay,
with
the
lowest
effective
concentration
for
TCAN
beginning
at
0.1
µ
g/
mL,
for
DBAN,
and
BCAN
at
0.12
µ
g/
mL,
and
for
DCAN
at
0.25
µ
g/
mL.
Although
all
of
the
compounds
significantly
increased
micronuclei
formation,
the
magnitude
of
this
effect
was
characterized
as
relatively
weak
for
DCAN
and
BCAN,
since
median
increases
in
the
number
of
micronucleated
erythrocytes
was
on
the
order
of
2­
fold
compared
to
controls.
Based
on
results
provided
in
a
summary
table,
the
maximum
increase
in
the
number
of
micronucleated
erythrocytes
for
TCAN
was
also
near
2­
fold,
while
for
DBAN
increases
were
larger
(
maximum
of
6.17­
fold).

In
another
assay
for
chromosome
damage,
DCAN
at
an
inhalation
concentration
of
8.6
ppm
induced
aneuploidy
in
the
offspring
of
female
Drosophila
melanogaster
exposed
for
up
to
45
minutes
(
Osgood
and
Sterling,
1991).
DBAN
was
highly
toxic;
a
concentration
of
only
0.3
ppm
killed
30­
40%
of
flies
after
a
45­
minute
exposure,
compared
to
8.6
ppm
for
DCAN.

However,
at
a
dose
of
0.3
ppm,
DBAN
did
not
induce
aneuploidy.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
54
EPA/
OW/
OST/
HECD
Final
Draft
A
variety
of
other
studies
have
been
conducted
to
test
whether
HANs
can
result
in
DNA
damage.
Bull
et
al.
(
1985)
studied
the
ability
of
HANs
to
induce
sister
chromatid
exchanges
(
SCE),
which
measures
a
repair
response
to
DNA
damage.
Chinese
hamster
ovary
(
CHO)
cells
were
treated
with
BCAN,
DBAN,
DCAN,
or
TCAN
(
concentrations
indicated
in
Table
V­
11)

without
exogenous
metabolic
activation
for
2
hours,
at
which
time
5­
bromo­
2­
deoxyuridine
(
BrdU)
was
added
to
the
medium
and
the
incubation
was
continued
for
26
to
32
hours.
The
cells
with
metabolic
activation
were
treated
with
HANs
for
2
hours
in
the
presence
of
S9,
at
which
time
the
cells
were
rinsed
and
fresh
medium
containing
BrdU
was
added
and
the
incubation
was
continued
for
26
hours.
The
average
SCE
frequency
was
significantly
elevated
in
the
presence
of
nonactivated
or
S9­
activated
for
BCAN,
DBAN,
DCAN,
and
TCAN.
Comparisons
of
the
potency
of
the
HANs
established
the
following:
DBAN
>
BCAN
>
TCAN
>
DCAN.

Zimmermann
et
al.
(
1984)
compiled
and
reviewed
published
results
of
genetic
damage
assays
in
Saccharomyces
cerevisiae
(
yeast).
The
only
HAN
compound
for
which
data
were
presented
was
DCAN
(
based
on
the
study
of
Simmon
et
al.,
1977;
and
Simmon
and
Kauhanen,

1978).
Based
on
their
analysis
of
the
original
report,
Zimmerman
and
colleagues
summarized
the
results
of
a
homozygosity
assay,
which
serves
to
indicate
gene
recombination
and
gene
conversion
events.
The
assay
was
conducted
in
a
stationary
culture
of
S.
cerevesiae
strain
D3
with
and
without
metabolic
activation
with
liver
microsomes.
DCAN
yielded
a
positive
result
in
this
assay
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
55
EPA/
OW/
OST/
HECD
Final
Draft
Table
V­
11.
Induction
of
Sister
Chromatid
Exchange
in
Chinese
Hamster
Ovary
Cells
by
BCAN,
DBAN,
DCAN,
and
TCAN.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
56
EPA/
OW/
OST/
HECD
Final
Draft
regardless
of
metabolic
activation.
In
a
more
complex
assay
system
in
yeast,
Zimmermann
and
Mohr
(
1992)
studied
the
effects
of
several
agents,
including
DBAN,
on
mitotic
chromosome
loss
and
mitotic
recombination.
Diploid
S.
cerevisiae
strain
D61.
M,
heterozygous
for
three
recessive
alleles
(
cyh2,
leu1,
and
ade6)
on
chromosome
VII,
was
used
to
test
for
effects
of
DBAN
on
chromosomal
malsegregation
or
mitotic
recombination.
Chromosome
loss
was
scored
based
on
the
number
of
colonies
expressing
all
three
recessive
markers,
and
mitotic
recombination
was
evaluated
based
on
the
expression
of
the
chy2
and
ade6
(
but
not
leu1
which
was
located
between
these
two
markers
on
the
chromosome).
Gene
expression
was
identified
by
colony
formation
on
the
appropriate
selective
plates.
The
yeast
cultures
were
treated
with
DBAN
at
concentrations
ranging
from
0
to
18.2
mg/
mL.
Mitotic
recombination
was
induced
in
a
dose­
dependent
fashion.

Chromosome
loss
was
not
the
reason
for
the
expression
of
the
recessive
markers.
In
contrast
to
this
result,
when
yeast
were
treated
with
DBAN
in
combination
with
propionitrile,
which
is
known
to
induce
chromosome
loss
and
enhances
the
sensitivity
of
the
malsegregation
analysis,
the
expected
loss
of
chromosomes
was
observed.
The
authors
speculated
that
failure
to
induce
malsegregation
with
DBAN
treatment
alone
reflects
high
toxicity
at
doses
that
would
induce
malsegregation.

The
studies
that
have
evaluated
DNA
damage
at
the
chromosome
level
have
resulted
in
inconsistent
results.
No
increase
in
micronuclei
was
reported
for
any
of
the
HANs
in
a
standard
assay
for
this
end
point
(
Bull
et
al.,
1985),
but
it
is
unclear
whether
high
enough
doses
were
tested.
Positive
results
were
reported
for
all
four
compounds
in
newt
larvae,
but
this
is
not
a
wellcharacterized
assay
system.
Positive
results
have
generally
been
observed
in
assays
that
measure
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
57
EPA/
OW/
OST/
HECD
Final
Draft
responses
to
DNA
damage,
including
sister
chromatid
exchange
in
CHO
cells
(
Bull
et
al.,
1985)

and
recombination
studies
in
yeast
(
Zimmermann
et
al.,
1984;
Zimmermann
and
Mohr,
1992).

Several
studies
have
been
conducted
to
evaluate
DNA
damage.
Le
Curieux
et
al.
(
1995)

studied
the
genotoxicity
of
BCAN,
DBAN,
DCAN,
and
TCAN
in
the
SOS
chromotest.
This
assay
measures
genotoxic
activity,
based
on
induction
of
the
SOS
DNA
repair
system
(
measured
by
increased
 ­
galactosidase
activity)
in
the
Escherichia
coli
strain
PQ37.
The
test
was
conducted
with
and
without
S9
activation.
The
range
of
doses
tested,
in
µ
g/
mL,
was:
BCAN,
3­

3000
(+/­
S9);
DBAN,
3­
1000
(+/­
S9);
DCAN
3­
1000
(+/­
S9);
TCAN,
3­
1000
(+
S9),
0.01­
1000
(­
S9).
TCAN
was
negative
up
to
cytotoxic
concentrations.
BCAN
and
DBAN
were
active
beginning
at
5
µ
g/
mL
and
10
µ
g/
mL,
respectively,
in
the
absence
of
S9,
but
were
inactive
in
the
presence
of
S9
activation.
DCAN
generated
a
positive
result
in
the
presence
of
S9
beginning
at
50
µ
g/
mL,
and
was
negative
in
the
absence
of
S9.
The
results
were
dose­
dependent
at
the
lower
doses,
but
decreased
at
the
higher
concentrations,
as
the
doses
exceeded
the
threshold
concentrations
for
cytotoxicity.
The
responses
seen
with
BCAN,
DBAN,
and
DCAN
were
considered
weak,
based
on
limited
induction
of
 ­
galactosidase
activity.

Lin
and
colleagues
in
a
series
of
papers
have
reported
on
the
ability
of
HANs
to
induce
direct
DNA
damage
by
measuring
DNA
strand
breaks,
the
ability
to
bind
to
the
nucleophilic
agent
(
4­
p­
nitrobenzyl­
pyridine),
and
formation
of
covalent
DNA
adducts
(
reviewed
in
Lin
et
al.,
1986).

Daniel
et
al.
(
1986)
reported
that
HANs
produced
DNA
strand
breaks
in
cultured
human
lymphoblastic
cells.
The
most
potent
initiator
of
DNA
strand
breaks
compared
to
control
cells
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
58
EPA/
OW/
OST/
HECD
Final
Draft
was
TCAN
which
induced
2­
fold
more
strand
breaks
than
the
positive
controls.
BCAN
and
DBAN
had
intermediate
activity,
while
DCAN
was
described
as
having
marginal
activity.
The
protocol
description
did
not
clearly
state
the
concentration
of
the
HANs
at
which
strand
breaks
were
observed,
but
exposure
to
50
µ
M
for
1
hour
resulted
in
cell
survival
decreases
ranging
from
40%
to
80%
of
control
cells.
The
HANs
also
showed
highly
variable
alkylation
potential
as
measured
by
the
potential
to
bind
to
4­
p­
nitrobenzyl­
pyridine
(
Daniel
et
al.,
1986).
The
relative
alkylation
potential
of
the
HANs
was
DBAN>
BCAN>>
DCAN>
TCAN.
The
range
in
reactivity
toward
4­
p­
nitrobenzyl­
pyridine
varied
by
627­
fold
among
the
HANs.

Daniel
et
al.
(
1986)
also
reported
the
results
of
DNA
adduct
analysis.
In
a
cell­
free
system,
[
14C]­
DCAN
was
incubated
with
calf
thymus
DNA,
the
DNA
was
precipitated,

hydrolyzed
and
separated
by
HPLC.
Similar
elution
peaks
were
observed
between
the
hydrolyzed
calf
thymus
DNA
and
incubations
of
14[
C]­
DCAN
with
polyadenylic
acid
or
polyguanylic
acid,

suggesting
that
DCAN
forms
an
adduct
with
these
nucleotides.
Oral
administration
of
DCAN
or
DBAN
to
rats
did
not
result
in
detectable
adduct
formation
in
liver
DNA
(
no
supporting
data
were
presented)
(
Lin
et
al.,
1986).
In
a
subsequent
study,
Lin
et
al.
(
1992)
reported
the
formation
of
DNA
adducts
following
gavage
dosing
of
radiolabeled
TCAN
to
male
F344
rats.
A
single
oral
gavage
dose
of
either
[
1­
14C]­
or
[
2­
14C]­
TCAN
(
in
tricaprylin,
7.2
 
69.3
mg/
kg)
was
administered
to
male
F344
rats
and
tissues
were
analyzed
from
2
to
48
hours
following
dosing.

TCAN
bound
to
both
DNA
and
proteins
and
DNA
binding
was
highest
in
the
stomach,
followed
by
liver
and
kidney.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
59
EPA/
OW/
OST/
HECD
Final
Draft
Overall,
the
data
suggest
that
HANs
can
directly
damage
DNA
as
evaluated
by
a
wide
array
of
assays
(
summarized
in
Table
V­
12).
The
weight
of
the
evidence
varies
for
each
compound.
BCAN
has
yielded
positive
results
in
all
assays
tested.
DBAN
yielded
negative
results
in
S.
typhimurium
mutation
assays
and
failed
to
form
DNA
adducts
in
vivo.
DBAN
appears
to
induce
DNA
strand
breaks
and
yield
positive
results
in
assays
that
reflect
responses
to
DNA
damage
(
i.
e.
SCE,
gene
conversion,
and
SOS
assays).
The
results
for
DCAN
and
TCAN
are
less
consistent.
DCAN
yielded
positive
results
in
S.
typhimurium
mutation
assays
and
assays
reflecting
DNA
recombination,
but
the
reason
for
absence
of
significant
effects
in
the
DNA
strand
break
assay
is
not
clear.
For
TCAN,
the
weak
responses
in
S.
typhimurium
mutation
assays
did
not
correspond
well
with
the
observed
in
vivo
formation
of
DNA
adducts,
although
the
positive
results
in
the
DNA
strand
break
assay
and
SCE
assay
were
consistent.

Table
V­
12.
Summary
of
Results
of
Genotoxicity
and
Tumor
Screening
Assays
for
Haloacetonitriles.

BCAN
DBAN
DCAN
TCAN
Mutation
assays
(
S.
typhimurium)
+
­
+
±
Micronuclei
±
±
±
±
Aneuploidy
(
D.
Melanogaster)
nt
­
+
nt
Sister
Chromatid
Exchange
+
+
+
+

Gene
Conversion/
Recombination
(
S.
cerevisiae)
nt
+
+
nt
SOS
chromotest
+
+
+
­

DNA
Strand
Breaks
+
+
­
+

DNA
adducts
(
in
vivo)
nt
­
­
+

Lung
tumors
(
A/
J
mice)
+
­
­
+

Skin
tumors
(
Sencar
mice)
+
+
­
­

nt
=
not
tested
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
60
EPA/
OW/
OST/
HECD
Final
Draft
The
evidence
for
the
induction
of
chromosome
damage
by
HANs
is
less
compelling,
due
to
the
limited
number
of
studies
available
for
evaluation
and
the
inconsistent
results.
In
the
single
study
that
used
a
standard
assay
protocol
to
evaluate
induction
of
micronuclei,
no
effect
was
observed
for
any
of
the
HANs,
although
is
was
not
clear
that
sufficiently
high
doses
were
used.

In
contrast,
positive
results
for
micronuclei
formation
were
reported
for
all
four
compounds
in
a
less
well
characterized
newt
larvae
system.
DCAN,
but
not
DBAN,
induced
aneuploidy
in
the
Drosophila
melanogaster
assay
system.

E.
Carcinogenicity
No
2­
year
carcinogenicity
bioassays
have
been
conducted
for
any
of
the
HANs
by
any
route
of
exposure.
No
alternative
carcinogenicity
studies
were
identified
for
any
the
HANs
by
the
inhalation
route.
There
are,
however,
several
short­
term
assays
that
can
aid
in
hazard
identification.
In
addition,
DBAN
is
currently
under
test
in
a
full
cancer
bioassay
by
NTP
(
2002).

In
a
published
review
paper,
Bull
and
Robinson
(
1985)
reported
studies
on
the
incidence
of
lung
tumors
in
groups
of
40
female
A/
J
mice
(
10
weeks
of
age)
that
were
administered
a
single
oral
dose
of
10
mg/
kg
of
BCAN,
DBAN,
DCAN,
or
TCAN,
three
times
weekly
for
8
weeks
(
Table
V­
13).
Control
groups
received
the
vehicle
only
(
10%
emulphor)
or
ethyl
carbamate
(
positive
control).
As
discussed
in
Chapter
VII,
emulphor
solutions
have
generally
been
deemed
as
more
appropriate
solvent
vehicles
than
corn
oil
for
disinfectant
byproducts.
All
animals
were
sacrificed
at
9
months
of
age,
allowing
for
an
approximately
6
months
post­
exposure
observation
period.
The
incidence
of
lung
tumors
(
adenomas)
was
significantly
increased
in
groups
given
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
61
EPA/
OW/
OST/
HECD
Final
Draft
BCAN
and
TCAN
(
p

0.05).
DBAN
and
DCAN
produced
marginal,
but
nonsignificant
(
p
>

0.05)
increases
in
lung
tumors.
The
authors
stated
that
the
results
should
be
interpreted
with
caution,
since
there
is
a
relatively
large
variation
in
the
background
incidence
of
lung
tumors
in
this
strain
of
mice
and
the
10
mg/
kg
dose
level
was
considerably
below
the
maximum
tolerated
dose,
decreasing
the
reliability
of
the
negative
findings
with
DBAN
and
DCAN.

Table
V­
13.
Effects
of
Orally
Administered
Haloacetonitriles
on
the
Development
of
Lung
Adenomas
in
Female
A/
J
Mice.

Number
of
%
Animals
Tumors/
Chemical
Dosea
animals
necropsied
w/
tumors
animal
Vehicle
(
emulphor
10%)
0.2
mL/
mouse
x24
31
10
0.10
BCAN
10
mg/
kg
x24
32
31
0.34b
DBAN
10
mg/
kg
x
24
31
16
0.19
DCAN
10
mg/
kg
x
24
30
23
0.23
TCAN
10
mg/
kg
x
24
32
28
0.38b
Ethyl
carbamate
(
positive
control)
42
mg/
kg
x
24
29
100
9.00
a.
Forty
female
strain
A/
J
mice
were
administered
the
indicated
doses
of
each
chemical
three
times
weekly
for
a
period
of
8
weeks.
Treatment
was
begun
at
10
weeks
of
age.
Animals
were
sacrificed
at
nine
months
of
age.

b.
Significantly
increased
above
controls
at
P<
0.05.

Adapted
from
Bull
and
Robinson
(
1985).

Bull
et
al.
(
1985)
studied
the
ability
of
BCAN,
DBAN,
DCAN,
and
TCAN
to
induce
tumors
in
mouse
skin
the
ability
of
dermally­
applied
HANs
to
act
as
tumor
initiators
was
studied
using
a
tumor
initiation/
promotion
protocol.
Six
topical
doses
of
0,
200,
400,
or
800
mg/
kg
HAN
dissolved
in
acetone
were
applied
to
the
shaved
backs
of
female
Sencar
mice
(
40
animals/
dose
group)
over
a
two­
week
period,
for
total
doses
of
0,
1200,
2400
and
4800
mg/
kg.
Beginning
two
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
62
EPA/
OW/
OST/
HECD
Final
Draft
weeks
after
the
last
HAN
dose,
1.0
µ
g
of
12­
O­
tetradecanoylphorbol­
13­
acetate
(
TPA)
was
applied
three
times
per
week
for
20
weeks.
Papilloma
incidence
and
regression
were
recorded
on
a
weekly
basis.
Animals
were
then
maintained
for
1
year
and
sacrificed
to
determine
the
incidence
of
squamous
cell
carcinomas.
The
results
from
this
initiation/
promotion
study
are
presented
in
Table
V­
14.
These
data
were
compiled
by
the
study
author
from
three
independent
experiments
as
indicated
in
the
column
of
the
table
labeled
"
Experiment
No".

Table
V­
14.
Histopathological
Diagnosis
of
Tumors
Resulting
from
Topical
Treatment
with
Halogenated
Haloacetonitriles
in
the
Sencar
Mouse
Treatment
Experiment
No.
Total
Dosea
(
mg/
kg)
Nb
Number
of
animals
with
squamous
cell
tumors
Squamous
cell
tumors
diagnosed
%
Animals
bearing
carcinomas
Number
%
Papilloma
Carcinoma
Acetone
1
2
3
0.2
ml
X
6
0.2
ml
X
6
0.2
ml
X
6
34
37
34
1
3
5
3
8
15
0
3
1
1
0
4
2.9
0
11.7
BCAN
2
2
2
1200
2400
4800
35
37
37
1
7
8
3
19
22*
0
0
3
1
7
6
2.9
18.9*
16.2*

DBAN
1
2
3
2
3
1200
2400
2400
4800
4800
36
35
35
37
37
8
17
16
7
3
22*
49**
46**
19
11
6
9
7
5
1
2
8
9
2
2
5.6
22.9**
25.7**
5.4
7.4
DCAN
1
2
2
1200
2400
4800
39
35
35
4
4
1
10
11
3
4
2
1
0
3
0
0
8.6
0
TCAN
1
2
3
2
3
1200
2400
2400
4800
4800
34
36
38
36
29
2
11
1
3
2
6
31**
3
8
7
1
5
0
2
1
1
6
1
1
1
2.9
16.7*
2.6
2.8
3.4
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
63
EPA/
OW/
OST/
HECD
Final
Draft
a
Total
doses
of
haloacetonitrile
shown
were
delivered
topically
in
six
equal
doses
over
a
2­
week
period.
Then
2
weeks
later
animals
were
treated
with
1.0
µ
g
12­
O­
tetradecanoylphorbol­
13­
acetate
(
TPA)
in
1.2
mL
acetone
topically
3
times
weekly
for
20
weeks.
b
There
were
40
animals
initiated
in
each
group,
N
=
number
available
for
histological
examination.
*
Significantly
different
from
the
control,
p
<
0.05,
Fisher
exact
test.
**
Significantly
different
from
control,
p
<
0.01,
Fisher
exact
test.

Adapted
from
Bull
et
al.
(
1985).

Both
BCAN
and
DBAN
induced
dose­
related
increases
in
the
percent
of
animals
with
squamous
cell
tumors
(
i.
e.,
combined
papillomas
and
carcinomas)
(
during
the
first
24
weeks
after
TPA
application
was
started)
and
the
percent
of
animals
with
carcinomas
(
at
1.5
years)(
Table
V­

14).
The
study
authors
indicated
that
the
decreased
tumor
response
for
DBAN
at
the
high
dose
as
compared
to
the
low
and
mid­
doses
may
have
resulted
from
the
severe
irritation
and
ulcerations
induced
by
this
treatment.
For
TCAN,
an
increase
in
the
percent
of
animals
with
squamous
cell
tumors
and
in
carcinomas
was
observed
at
the
mid­
dose
(
in
experiment
2),
but
this
increase
was
not
replicated
in
experiment
3
at
this
dose.
The
combined
data
for
TCAN
did
not
yield
a
significant
increase
in
tumor
response.
No
significant
increase
in
papillomas
or
carcinomas
was
observed
with
DCAN.
A
similar
pattern
of
results
for
each
of
the
HANs
was
obtained
when
the
number
of
tumors/
animal
was
evaluated
as
the
response
metric.
Based
on
these
results,
the
authors
concluded
that
DBAN
is
the
most
potent
mouse
skin
tumor
initiator
of
the
HANs
tested.

DBAN
is
followed
in
potency
by
BCAN,
while
TCAN
and
DCAN
were
judged
as
ineffective
inducers
of
skin
tumors
in
this
screening
bioassay.

Bull
et
al.
(
1985)
conducted
a
second
series
of
studies,
designed
to
assess
the
ability
of
orally­
administered
HANs
to
act
as
tumor
initiators.
In
this
study,
total
oral
doses
of
50
mg/
kg
were
administered
to
female
Sencar
mice
six
times
over
a
2­
week
period.
The
promotion
phase
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
64
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OW/
OST/
HECD
Final
Draft
of
the
study
was
conducted
using
the
same
procedures
as
for
the
dermal­
dosing
initiation/
promotion
study
described
above.
No
statistically­
significant
increase
in
tumor
yield
or
decrease
in
the
time­
to­
tumor
was
observed
for
any
of
the
HANs
when
the
data
across
all
oral
experiments
were
combined.
Sporadic
increases
in
tumor
yield
at
various
times
for
individual
HANs
were
noted
(
data
not
shown
by
the
authors),
but
were
not
judged
by
the
study
authors
to
be
biologically
significant.
The
positive
control,
300
mg/
kg
urethane,
increased
the
yield
of
papillomas.

In
a
third
cancer
screening
study
presented
in
the
paper
by
Bull
et
al.
(
1985),
the
ability
of
dermally­
applied
HANs
to
act
as
complete
carcinogens
was
assessed.
For
this
study,
BCAN,

DCAN,
or
TCAN
(
800
mg/
kg),
or
DBAN
(
400
mg/
kg)
was
applied
to
the
skin
of
female
Sencar
mice
3
times
per
week
for
24
weeks,
and
the
number
of
squamous
cell
tumors
recorded.
None
of
the
HANs
induced
skin
tumors
in
this
assay
(
data
not
shown
by
the
authors).

DBAN,
DCAN,
and
TCAN
were
tested
for
initiating
activity
using
the
rat
liver
gamma­
glutamyltranspeptidase
foci
(
GGT­
foci)
assay
in
F344
rats
as
an
indicator
of
carcinogenicity
(
Herren­
Freund
and
Pereira,
1986;
Lin
et
al.,
1986).
The
protocol
for
this
assay
consists
of
a
two­
thirds
partial
hepatectomy
followed
18
or
24
hours
later
by
the
administration
of
the
initiator.
The
doses
employed
were
2.0
mmol/
kg
(
398
mg/
kg)
for
DBAN,
2.0
mmol/
kg
(
220
mg/
kg)
for
DCAN,
and
1.0
mmol/
kg
(
144
mg/
kg)
for
TCAN.
Seven
days
after
initiation,
the
rats
received
the
promoter
(
500
ppm
sodium
phenobarbital
in
their
drinking
water)
for
at
least
ten
weeks.
The
halogenated
acetonitriles
were
inactive
as
initiators
in
the
GGT­
foci
assay.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
65
EPA/
OW/
OST/
HECD
Final
Draft
The
existing
data
provide
at
best
only
marginal
support
for
the
conclusion
that
HANs
are
carcinogenic.
The
evidence
is
stronger
for
BCAN,
which
increased
tumor
yields
in
both
lung
tumor
and
dermal
screening
assays.
DBAN
was
positive
at
non­
ulcerative
doses
in
the
dermal
screening
assay.
TCAN
was
positive
only
in
the
lung
assay,
and
DCAN
treatment
did
not
increase
either
lung
or
skin
tumors.
Opposing
these
positive
findings
are
the
negative
results
for
DBAN,
DCAN,
and
TCAN
in
the
GGT
foci
assay.
Overall,
the
data
are
insufficient
to
qualitatively
or
quantitatively
assess
the
carcinogenic
potential
of
any
of
the
HANs.
The
positive
results
in
two
tumor
screening
assays,
together
with
positive
bacterial
gene
mutation
results,

suggest
that
it
would
be
worthwhile
to
conduct
a
full
2­
year
bioassay
for
BCAN.
Results
for
the
other
HANs
are
mixed,
with
inconsistencies
between
the
genotoxicity
and
tumor
screening
data.

F.
Summary
The
toxicity
data
on
the
HANs
are
summarized
in
Tables
V­
15
through
V­
18.
Overall,

very
little
data
are
available
evaluating
the
non­
cancer
effects
of
the
HANs.
Acute
oral
LD
50
values
for
DBAN,
DCAN,
and
TCAN
in
rodents
have
been
reported
to
range
from
50
to
361
mg/
kg.
DBAN
and
TCAN
have
been
reported
to
be
irritants.
DBAN
causes
eye,
nasal,
and
respiratory
tract
irritation
following
inhalation,
and
skin
irritation
following
dermal
exposure.

TCAN
also
causes
skin
irritation
following
dermal
exposure.
No
data
on
the
acute
toxicity
of
BCAN
are
available,
and
no
subacute
or
subchronic
studies
of
either
BCAN
or
TCAN
are
available.
No
chronic
studies
have
been
conducted
on
any
of
the
HANs.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
66
EPA/
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OST/
HECD
Final
Draft
No
target
organ
has
been
clearly
established
for
HANs
following
oral
exposure,
although
absolute
and
relative
organ
weight
changes,
including
decreased
testes
weight
(
NTP,
2002)
and
increased
liver
weight
(
Hayes
et
al.,
1986;
Christ
et
al.,
1996)
have
been
reported.
Fourteen­
day
or
longer
systemic
toxicity
studies
have
been
conducted
in
mice
and
rats.
In
14­
day
and
90­
day
studies
of
DBAN
(
NTP,
2002;
Hayes
et
al.,
1986),
consistent,
compound­
related,

dosedependent
effects
were
limited
to
decreased
water
consumption,
decreased
body
weight,
and
decreased
testes
weight
and
pathology.
However,
effects
on
the
testes
reported
in
the
NTP
(
2002)
study
were
observed
only
in
rats
in
the
14­
day
study.
No
effects
on
the
testes
were
observed
in
rats
in
the
13­
week
study,
or
in
mice
(
NTP,
2002).
In
addition,
no
effects
were
observed
on
the
testes
in
rats
in
a
14­
day
or
90­
day
gavage
study
in
rats,
even
at
much
higher
doses
(
Hayes
et
al.,
1986).
For
DBAN,
the
observed
liver
weight
increases
reported
in
Hayes
et
al.
(
1986)
were
not
supported
by
other
measures
of
liver
toxicity
in
the
same
study
or
observed
in
the
more
recent
NTP
study
(
NTP,
2002),
and
therefore,
this
endpoint
was
not
selected
as
the
basis
for
the
quantitative
dose­
response
assessment.
Taken
together,
the
data
suggest
that
decreased
body
weight
appears
to
be
the
primary
indicator
of
toxicity
for
DBAN.
Overall,
male
rats
appear
to
be
more
sensitive
than
female
rats
for
DBAN.
For
DCAN,
consistent,

compoundrelated
dose­
dependent
effects
were
limited
to
decreased
body
weight
and
increased
liver
weight
(
Hayes
et
al.,
1986).
In
this
case,
the
observed
liver
weight
increases
were
supported
by
changes
in
serum
biochemistry
parameters
suggestive
of
liver
damage.
No
histopathological
evaluation
was
done
in
the
key
study
for
DCAN
(
Hayes
et
al.,
1986),
so
the
degree,
if
any,
of
liver
damage
can
not
be
confirmed.
The
data
for
TCAN
and
BCAN
are
too
limited
to
identify
with
confidence
any
potential
target
organs.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
67
EPA/
OW/
OST/
HECD
Final
Draft
The
data
are
inadequate
to
determine
whether
HANs
are
reproductive
toxicants.
No
multigeneration
reproductive
toxicity
study
has
been
conducted.
BCAN,
DBAN,
DCAN,
and
TCAN
at
doses
of
up
to
50
mg/
kg/
day
had
no
effect
on
sperm
morphology
(
Meier
et
al.,
1985),

but
the
data
on
testes
weight
changes
are
mixed
(
NTP,
2002;
Hayes
et
al.,
1986).
DBAN
at
doses
up
to
approximately
10
mg/
kg/
day
had
no
effect
on
any
male
or
female
reproductive
parameter
evaluated
in
a
screening
assay
(
R.
O.
W
Sciences,
1997).
A
series
of
developmental
toxicity
studies
in
rats
has
also
been
conducted.
Exposure
to
BCAN
and
DBAN
on
gestation
days
7
to
21
resulted
in
reduced
mean
birth
weight
(
Smith
et
al.,
1986;
Smith
et
al.,
1987).
It
addition
to
this
effect,
DCAN
and
TCAN
decreased
the
percentage
of
females
delivering
viable
litters
and
increased
fetal
resorptions
(
Smith
et
al.,
1986;
Smith
et
al.,
1987;
Smith
et
al.,
1988;

Smith
et
al.,
1989;
Christ
et
al.,
1996).
DCAN
and
TCAN
also
significantly
increase
the
frequency
of
malformations
in
fetuses
(
Smith
et
al.,
1988;
Smith
et
al.,
1989;
Christ
et
al.,
1996).

These
studies
by
Smith
and
colleagues
on
the
developmental
toxicity
of
HANs
in
rats
were
conducted
using
tricaprylin
as
a
vehicle,
because
these
compounds
are
very
miscible
in
this
vehicle.
However,
in
these
studies,
comparison
of
tricaprylin
versus
water­
treated
controls
revealed
increased
embryotoxicity
due
to
tricaprylin.
A
recent
study
by
Christ
et
al.
(
1996)

indicates
that
tricaprylin
also
influences
the
pattern
of
malformations
observed
in
fetuses
caused
by
TCAN.
For
TCAN
in
corn
oil,
the
malformations
were
primarily
cranio­
facial
in
nature
while
for
TCAN
in
tricaprylin
the
malformations
were
primarily
cardiovascular
and
urogenital
in
nature.

Therefore,
the
use
of
data
from
studies
in
which
tricaprylin
was
used
as
the
vehicle
is
not
appropriate
for
risk
assessment
purposes.
In
the
one
developmental
toxicity
study
that
used
a
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
68
EPA/
OW/
OST/
HECD
Final
Draft
vehicle
other
than
tricaprylin
(
Christ
et
al.,
1996),
maternal
toxicity
was
observed
at
lower
doses
than
developmental
effects.

Overall,
the
data
suggest
that
HANs
can
directly
damage
DNA
as
evaluated
by
a
wide
array
of
assays
(
summarized
in
Table
V­
12).
The
weight
of
the
evidence
varies
for
the
each
compound.
BCAN
has
yielded
positive
results
in
all
assays
tested.
DBAN
yielded
negative
results
in
S.
typhimurium
mutation
assays
and
failed
to
form
DNA
adducts
in
vivo.
DBAN
appears
to
induce
DNA
strand
breaks
and
yield
positive
results
in
assays
that
reflect
responses
to
DNA
damage
(
i.
e.
SCE,
gene
conversion,
and
SOS
assays).
The
results
for
DCAN
and
TCAN
are
less
consistent.
DCAN
yielded
positive
results
in
S.
typhimurium
mutation
assays
and
assays
reflecting
DNA
recombination,
but
the
reason
for
absence
of
significant
effects
in
the
DNA
strand
break
assay
is
not
clear.
For
TCAN,
the
weak
responses
in
S.
typhimurium
mutation
assays
did
not
correspond
well
with
the
observed
in
vivo
formation
of
DNA
adducts,
although
the
positive
results
in
the
DNA
strand
break
assays
and
SCE
assay
were
consistent.

The
evidence
for
the
induction
of
chromosome
damage
by
HANs
is
less
compelling,
due
to
the
limited
number
of
studies
available
for
evaluation
and
the
inconsistent
results.
In
the
single
study
that
used
a
standard
assay
protocol
to
evaluate
induction
of
micronuclei,
no
effect
was
observed
for
any
of
the
HANs,
although
is
was
not
clear
that
sufficiently
high
doses
were
tested.

In
contrast,
positive
results
for
micronuclei
formation
were
reported
for
all
four
compounds
in
a
less
well
characterized
newt
larvae
system.
DCAN,
but
not
DBAN
induced
aneuploidy
in
Drosophila
melanogaster
assay
system.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
69
EPA/
OW/
OST/
HECD
Final
Draft
The
existing
data
provide
at
best
only
marginal
support
for
the
conclusion
that
HANs
are
carcinogenic.
The
evidence
is
stronger
for
BCAN,
which
increased
tumor
yields
in
both
lung
tumor
and
dermal
screening
assays
(
Bull
and
Robinson,
1985;
Bull
et
al.,
1985).
DBAN
was
positive
at
non­
ulcerative
doses
in
the
dermal
screening
assay.
TCAN
was
positive
only
in
the
lung
assay,
and
DCAN
treatment
did
not
increase
either
lung
or
skin
tumors.
Opposing
these
positive
findings
are
the
negative
results
for
DBAN,
DCAN,
and
TCAN
in
the
GGT
foci
assay
(
Herren­
Freund
and
Pereira,
1986).
Overall,
the
data
are
insufficient
to
qualitatively
or
quantitatively
assess
the
carcinogenic
potential
of
any
of
the
HANs.
The
positive
results
in
two
tumor
screening
assays,
together
with
positive
bacterial
gene
mutation
results,
suggest
that
it
would
be
worthwhile
to
conduct
a
full
2­
year
bioassay
for
BCAN.
DBAN
is
currently
on
test
for
a
full
cancer
bioassay
(
NTP,
2002).
Results
for
the
other
HANs
are
more
mixed,
with
inconsistencies
between
the
genotoxicity
and
tumor
screening
data.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
70
EPA/
OW/
OST/
HECD
Final
Draft
Table
V­
15.
Summary
of
Oral
Studies
of
BCAN
Toxicity.

Reference
Species/
Strain
Route
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.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
71
EPA/
OW/
OST/
HECD
Final
Draft
Table
V­
16.
Summary
of
Oral
Studies
of
DBAN
Toxicity.

Reference
Species/
Strain
Route
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
Unspecified
Gavage
25
­
1,600
mg/
kg/
day
Acute
Lethality
ND
LD50
=
50
Rat
Unspecified
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
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

V
­
72
EPA/
OW/
OST/
HECD
Final
Draft
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
(
Freestanding
NOAEL)
Drinking
Water
Criteria
Document
for
Haloacetonitriles
Reference
Species/
Strain
Route
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

V
­
73
EPA/
OW/
OST/
HECD
Final
Draft
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)

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
periconcep
tion
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
/
development
al:
8.2
(
M);
10.8
(
F)

(
Freestanding
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
Development
al:
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
V
­
74
EPA/
OW/
OST/
HECD
Final
Draft
Table
V­
17.
Summary
of
Oral
Studies
of
DCAN
Toxicity.

Reference
Species/
Strain
Route
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
Exposure
Duration
Endpoints
Evaluated
NOAEL
(
mg/
kg/
day)
LOAEL
(
mg/
kg/
day)

V
­
75
EPA/
OW/
OST/
HECD
Final
Draft
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.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
76
EPA/
OW/
OST/
HECD
Final
Draft
Table
V­
18.
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.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
V
­
77
EPA/
OW/
OST/
HECD
Final
Draft
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.
