Document ID: EPA-HQ-OW-2002-0043-0205
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2003-08-11T04:00Z

Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
1
Final
Draft
Chapter
VII.
Mechanism
of
Toxicity
and
Sensitive
Subpopulations
A.
Biochemical
Basis
of
Toxicity
Both
DBAN
and
DCAN
are
general
systemic
toxicants,
with
depressed
body
weight
identified
as
a
common
adverse
effect
(
NTP,
2002;
Hayes
et
al.,
1996).
Although
other
organ
weight
changes
were
observed
at
doses
associated
with
decreased
body
weight,
the
liver
is
an
organ
for
which
dose­
dependent
effects
on
organ
weight
were
supported
by
other
measures
of
toxicity.
Increased
enzyme
levels
(
e.
g.,
ALP)
in
the
serum
were
observed
at
doses
higher
than
those
that
induced
liver
weight
changes
(
Hayes
et
al.,
1986),
suggesting
that
DBAN
and
DCAN
might
induce
hepatocellular
necrosis.
However,
ALP
is
not
a
liver
specific
enzyme,
and
changes
in
the
liver
specific
enzymes
SGPT
and
SGOT
were
not
consistently
dose­
related.
In
addition,

Hayes
et
al.
(
1986)
did
not
perform
histopathological
examinations,
and
therefore,
cellular
changes
resulting
in
increased
liver
weight
could
not
be
determined.
The
data
suggest
that
DCAN
may
be
the
more
potent
liver
toxicant
compared
to
DBAN,
since
the
observed
effects
in
the
gavage
study
(
Hayes
et
al.,
1986)
for
DCAN
were
more
severe
than
for
DBAN.
Furthermore,

DBAN
administered
in
drinking
water
did
not
induce
liver
toxicity,
other
than
a
treatment­
related
increase
in
liver
GST
activity
(
NTP,
2002).
However,
the
highest
doses
in
the
drinking
water
study
(
NTP,
2002)
were
similar
to
the
NOAEL
in
the
Hayes
et
al.
(
1986)
study.
None
of
the
available
data
are
adequate
to
determine
mechanisms
of
liver
toxicity
for
the
HANs.

One
postulated
mode
of
action
for
the
toxicity
of
haloacetonitriles
(
HANs)
is
through
direct
interactions
with
cellular
macromolecules
(
Pereira
et
al.,
1984;
Daniel
et
al.,
1986;
Lin
and
Guion,
1989;
Lin
et
al.,
1992).
Depletion
of
reduced
glutathione
(
GSH)
could
also
play
a
role.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
2
Final
Draft
HANs
have
been
shown
to
induce
transient
decreases
in
rat
liver
GSH
levels
and
inhibit
glutathione­
S­
transferase
activity
in
vivo
(
Lin
and
Guion,
1989;
Ahmed
et
al.,
1991).
Ahmed
et
al.
(
1989)
noted
that
the
relative
degree
of
inhibition
of
rat
liver
glutathione­
S­
transferase
(
GST)

activity
in
vitro,
reported
as
TCAN>
DBAN>
DCAN,
is
consistent
with
the
relative
toxicity
of
these
compounds
reported
in
the
literature,
suggesting
perturbation
of
GSH
protection
as
an
important
mechanism
of
toxicity.
As
further
support
for
the
relatedness
of
GSH
depletion
and
HAN­
induced
toxicity,
Ahmed
et
al.
(
1991)
noted
that
the
sustained
depletion
of
cellular
GSH
levels
in
stomach
tissues
by
DBAN
was
consistent
with
their
earlier
preliminary
finding
that
acute
oral
doses
of
HANs
can
damage
the
gastric
tissues.
However,
the
effect
of
HANs
on
gastric
tissue
might
simply
reflect
the
direct
irritancy
of
these
compounds,
rather
than
GSH
depletion.

The
initial
finding
of
HANs
damaging
gastric
tissues
cannot
be
further
investigated,
because
the
subacute
and
subchronic
studies
(
Hayes
et
al.,
1986)
did
not
include
a
histopathological
examination
of
the
stomach.

GSH
depletion
could
enhance
cytotoxicity
by
allowing
damage
to
cellular
macromolecules
by
HANs,
their
metabolites,
or
other
reactive
species
accumulated
in
the
cell.
Since
GSH
is
an
important
cellular
antioxidant,
its
depletion
might
induce
cellular
oxidative
stress.
In
support
of
this
idea,
Ahmed
et
al.
(
1999)
reported
that
orally­
administered
monochloroacetonitrile
(
MCAN)

induced
a
dose­
dependent
decrease
in
GSH
levels
and
increased
levels
of
oxidative
DNA
damage
in
the
stomach
mucosa
of
rats.
Alternative
mechanisms
for
MCAN­
induced
oxidative
stress
were
discussed
by
the
study
authors.
One
proposed
mechanism
involves
GSH
depletion,
resulting
in
decreased
ability
of
the
cell
to
detoxify
endogenous
reactive
oxygen
species.
In
an
alternative
mechanism,
cyanide
derived
from
MCAN
metabolism
might
alter
cellular
oxygen
utilization
and
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
3
Final
Draft
increase
the
formation
of
reactive
oxygen
species.
In
agreement
with
the
proposed
role
of
oxidative
stress,
Mohamadin
and
Abdel­
Naim
(
1999)
reported
that
MCAN
decreased
cellular
GSH
content
and
increased
lipid
peroxidation,
as
measured
by
the
concentration
of
thiobarbituric
acid
reactive
substances
(
TBARS),
in
rat
gastric
epithelial
cells
in
culture.
Cell
viability,
measured
by
the
release
of
lactate
dehydrogenase
activity,
was
correlated
with
the
depletion
of
cellular
GSH
levels
®
=
0.96).
Supplementing
the
culture
medium
with
treatments
that
protect
against
cellular
oxidative
stress
(
e.
g.
antioxidants
or
iron
chelators),
decreased
the
cytotoxicity
and
lipid
peroxidation
induced
by
MCAN.
MCAN
is
not
a
compound
under
review
for
this
document.

However,
since
it
shares
the
ability
to
deplete
GSH
and
form
cyanide
with
other
HANs,
these
data
are
appropriate
for
discussion
here,
although
in
generalizing
about
the
HANs
as
a
class
of
compounds,
differences
in
their
ability
to
deplete
GSH
should
be
considered
in
evaluating
the
likely
mechanisms
for
any
individual
compound.

In
addition
to
the
induction
of
oxidative
stress
secondary
to
GSH
depletion
or
cyanide
activity
on
cell
respiration,
an
alternative
pathway
might
include
activation
of
macrophages
to
release
reactive
oxygen
species.
Ahmed
et
al.
(
2000)
reported
that
DCAN
induced
oxidative
stress
responses,
including
increased
oxidation
of
glutathione,
increased
formation
of
reactive
oxygen
intermediates,
and
increased
tumor
necrosis
factor
alpha
(
TNF­
 )
secretion
(
a
cellular
response
during
macrophage
activation)
in
a
mouse
macrophage
cell
line
in
culture.
To
explain
these
data,
the
authors
proposed
that
DCAN
treatment
activates
macrophages,
with
subsequent
increases
in
reactive
oxygen
intermediate
production
and
in
TNF­
 
secretion.
The
increased
production
of
reactive
oxygen
species
induces
oxidative
stress
that
reduces
cell
viability
through
apoptosis
and
necrosis.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
4
Final
Draft
The
identification
of
thiocyanate
as
a
urinary
metabolite
in
animals
orally­
dosed
with
HANs,
and
the
hypothesized
metabolism
to
cyanide
(
Pereira
et
al.,
1984),
suggest
additional
possible
effects
of
HANs
that
should
be
investigated.
Longer­
term
exposure
to
thiocyanate
(
either
from
thiocyanate
dosing
or
as
a
metabolite
of
cyanide)
causes
thyroid
effects.
Central
nervous
system
effects
(
e.
g.,
myelin
degeneration)
and
male
reproductive
effects
(
decreased
epididymal
weight
and
sperm
motility)
have
also
been
observed
following
long­
term
exposure
to
cyanide
(
U.
S.
EPA,
2002c).
Effects
on
the
central
nervous
system
or
thyroid
were
not
observed
in
a
recent
14­
day
or
13­
week
NTP
(
2002)
study
for
DBAN.
Although
decreased
testes
weight
and
testes
atrophy
were
reported
in
the
14­
day
(
NTP,
2002)
study
for
DBAN,
this
effect
has
not
been
observed
in
subchronic
studies
(
Hayes
et
al.,
1986;
NTP,
2002),
or
in
other
studies
that
evaluated
male
reproductive
tract
parameters
(
R.
O.
W.
Sciences,
1997;
Meier
et
al.,
1985).

However,
only
limited
conclusions
can
drawn
regarding
the
potential
for
HANs
to
induce
a
similar
array
of
effects
as
cyanide
and
thiocyanate,
based
on
limitations
in
the
overall
database.

A
comparison
of
relative
potency
among
HANs
in
thiocyanate
excretion,
alkylation
potential
(
p­
nitrobenzopyridine
binding),
protein
binding
(
inhibition
of
dinitrosamine
demethylase
activity),
and
production
of
DNA
strand
breaks,
suggested
that
relative
thiocyanate
excretion
did
not
correspond
well
to
some
of
these
markers
of
potential
macromolecule
interaction
(
Pereira
et
al.,
1984).
For
example,
monochloroacetonitrile
(
MCAN)
was
the
most
potent
thiocyanate
former,
but
TCAN
was
more
potent
than
MCAN
in
protein
binding
and
inducing
DNA
strand
breaks.
The
authors
suggested
that
this
discordance
between
propensity
for
cyanide
formation
and
induction
of
toxic
outcomes
might
reflect
the
formation
of
reactive
intermediates
other
than
cyanide
from
TCAN,
such
as
phosgene
and
cyanoformyl
chloride.
The
results
of
the
comparative
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
5
Final
Draft
analysis
by
Pereira
et
al.
(
1984)
suggest
that
intermediate
metabolites,
other
than
cyanide,
may
be
important
in
producing
some
of
the
toxic
effects
observed
for
HANs.
In
further
support
of
this
conclusion,
the
systemic
toxicity
induced
by
cyanide
and
thiocyanate
does
not
closely
parallel
the
range
of
effects
observed
for
HANs.
For
example,
cyanide
and
thiocyanate
are
not
potent
liver
toxicants,
a
target
organ
for
the
effects
of
HANs
(
U.
S.
EPA,
2002c).

BCAN,
DBAN,
DCAN,
and
TCAN
all
cause
developmental
toxicity
in
vivo
(
Smith
et
al.,

1986;
Smith
et
al.,
1987;
Smith
et
al.,
1988;
Smith
et
al.,
1989;
Christ
et
al.,
1995;
Christ
et
al.,

1996),
but
the
relative
potency
of
these
compounds
is
unclear,
due
to
the
likely
potentiating
effects
of
the
solvent
vehicle
(
tricaprylin)
used
in
these
studies
(
Christ
et
al.,
1996).

As
discussed
in
Chapter
III
(
Toxicokinetics),
it
remains
unclear
whether
the
potentiating
effect
of
tricaprylin
on
the
developmental
toxicity
of
HANs
represents
a
toxicokinetic
or
toxicodynamic
interaction,
or
whether
both
of
these
mechanisms
play
a
role.
Since
the
data
on
tricaprylin
effects
for
HANs
were
limited
to
two
published
abstracts
(
Roth
et
al.,
1990;
Gordon
et
al.,
1991),
a
review
of
the
literature
on
interactions
between
solvent
vehicles
and
other
disinfectant
by­
products
or
unsaturated
nitriles
was
done
to
determine
whether
a
consistent
relationship
could
be
found.
Solvent
vehicles
are
needed
for
studies
with
these
compounds,
due
to
the
minimal
water
solubility.
No
data
were
found
for
interactions
between
tricaprylin
and
the
comparison
compounds.
Therefore,
data
on
observed
solvent
vehicle
interactions
are
too
limited
for
trihalomethanes
and
unsaturated
nitriles
to
be
useful
in
reaching
a
general
conclusion
about
the
mechanisms
involved
in
the
ability
of
tricaprylin
to
potentiate
the
developmental
toxicity
of
HANs.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
6
Final
Draft
Developmental
toxicity
has
been
observed
even
in
the
absence
of
confounding
by
tricaprylin,
at
least
for
TCAN
(
Christ
et
al.,
1996).
Maternal
toxicity
was
observed
at
lower
doses
than
developmental
effects
in
this
study,
and
therefore,
it
is
possible
that
the
observed
developmental
effects
were
secondary
to
maternal
effects.
Mechanistic
explanations
for
the
developmental
toxicity
of
HANs
have
been
explored,
including
cyanide
formation
and
glutathione
depletion.
Early
evidence
suggested
that
the
developmental
toxicity
of
HANs
might
not
be
secondary
to
cyanide
formation.
When
a
maximally
tolerated
dose
of
a
series
of
HANs
in
a
tricaprylin
vehicle
was
administered
to
Long­
Evans
rats,
developmental
toxicity
was
greatest
for
the
highly
chlorine­
substituted
acetonitriles
(
Smith
et
al.,
1986).
These
results
were
contrasted
with
the
work
of
Pereira
et
al.
(
1984),
who
found
that
thiocyanate
excretion
(
and
hence
cyanide
production)
is
inversely
related
to
chlorine
substitution.
Based
on
this
comparison,
in
which
greater
chlorine
substitution
increases
developmental
toxicity
and
decreases
thiocyanate
excretion,

the
authors
concluded
that
the
degree
of
developmental
toxicity
was
not
due
to
cyanide
formation.
However,
conclusions
drawn
from
the
HANs
studies
must
be
tempered
by
uncertainty
regarding
their
true
relative
developmental
toxicity
potencies.
The
tricaprylin
vehicle
used
in
the
developmental
toxicity
studies
causes
developmental
effects
by
itself
(
Smith
et
al.,
1989;
Christ
et
al.,
1995),
and
interacts
with
TCAN
to
cause
both
qualitative
changes
in
the
spectrum
of
developmental
effects,
as
well
as
more­
than­
additive
quantitative
changes.
Because
of
these
effects
of
the
vehicle
used
in
the
relevant
developmental
toxicity
studies,
and
because
the
relative
potentiating
ability
of
tricaprylin
may
differ
for
each
of
the
HANs
tested,
the
true
relative
developmental
toxicity
potencies
are
unknown.
Another
potential
avenue
for
determining
the
contribution
of
cyanide
to
the
developmental
toxicity
of
HANs
would
be
to
compare
the
developmental
effects
observed
in
animals
exposed
to
cyanide
versus
those
observed
for
HANs.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
7
Final
Draft
However,
the
available
developmental
toxicity
studies
for
both
HANs
(
see
Chapter
V)
and
cyanide
(
U.
S.
EPA,
2002c)
are
too
limited
to
conduct
a
meaningful
comparison
of
their
relative
potencies
as
developmental
toxicants.

Christ
et
al.
(
1995),
in
a
study
on
the
developmental
toxicity
of
BCAN,
reiterated
the
discordance
between
developmental
toxicity
and
degree
of
metabolism
to
cyanide,
and
introduced
GSH
depletion
as
another
possible
mechanism
for
developmental
toxicity.
To
address
the
role
of
GSH
depletion
in
the
developmental
toxicity
of
HANs,
Christ
et
al.
(
1995)
cited
the
study
of
Abdel­
Aziz
et
al.
(
1993),
who
compared
maternal
and
fetal
uptake
of
[
2­
14C]­
MCAN
in
CD­
1
mice.
Abdel­
Aziz
et
al.
(
1993)
treated
dams
with
diethylmaleate
(
DEM)
by
intraperitoneal
injection
to
induce
an
oxidative
stress
response
on
gestation
day
13.
One
hour
later,
control
mice
and
DEM­
pretreated
mice
were
given
a
77
mg/
kg
dose
of
radiolabeled
MCAN
by
i.
v.
injection.

MCAN
treatment
significantly
decreased
maternal
liver,
uterus,
and
fetal
tissue
levels
of
GSH,

and
the
degree
of
GSH
depletion
was
further
increased
in
the
DEM­
pretreated
mice.
Urinary
excretion
of
thiocyanate,
a
measure
of
MCAN
metabolism,
was
five
times
higher
in
the
DEMpretreated
mice
given
MCAN
as
compared
to
control
mice
treated
with
MCAN
only.
MCAN
equivalents
determined
from
the
yield
of
radioactivity
in
maternal
uterine
tissues,
amniotic
fluid,

and
in
fetuses
rose
rapidly
to
similar
levels
at
1
hour
for
both
DEM­
pretreated
mice
and
mice
not
given
DEM.
In
the
absence
of
DEM
pretreatment,
MCAN
equivalents
declined
rapidly
for
all
three
tissues
examined.
However,
in
DEM­
pretreated
animals,
the
removal
of
tissue
radioactivity
was
significantly
slower.
At
24
hours,
the
level
of
MCAN
equivalents
was
two­
fold
higher
in
fetal
DNA
than
in
maternal
uterine
DNA.
This
effect
was
further
enhanced
by
DEM
pretreatment.

The
total
DNA­
bound
MCAN
equivalent
level
in
DEM­
pretreated
mice
was
four­
fold
higher
in
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
8
Final
Draft
fetal
DNA
than
in
maternal
uterine
DNA,
suggesting
that
fetuses
might
be
particularly
sensitive
to
the
effects
of
GSH
depletion.
Based
on
the
increases
in
thiocyanate
excretion
and
DNA
binding
under
oxidative
stress
conditions
that
deplete
GSH,
the
authors
suggested
that
GSH
depletion
increases
the
metabolism
of
the
remaining
unconjugated
MCAN
and/
or
increases
the
availability
of
MCAN
to
react
directly
with
cellular
macromolecules
such
as
DNA.
These
data
suggest
that
the
developmental
toxicity
of
HANs
may
be
directly
related
to
the
ability
of
these
compounds
to
deplete
GSH
levels.

Saillenfait
and
Sabate
(
2000)
tested
the
developmental
toxicity
of
aliphatic
nitriles
(
sodium
cyanide,
acetonitrile,
propionitrile,
n­
butyronitrile,
acrylonitrile,
methacrylonitrile,
allylnitrile,
cis­

2­
pentenenitrile,
and
2­
chloroacrylonitrile)
in
a
rat
whole
embryo
assay
and
in
vivo.
Although
no
HANs
were
included
in
this
study,
the
results
may
have
implications
for
the
developmental
toxicity
of
HANs
based
on
similar
chemical
reactivity.
In
the
whole
embryo
testing
experiment,
a
wide
range
of
embryotoxicity
was
observed.
In
addition,
no
common
pattern
of
developmental
effects
was
observed
among
all
the
compounds
tested.
Enhancement
of
metabolism
by
supplementation
of
the
cultures
with
microsomes
increased
the
severity
of
embryolethality
and
developmental
toxicity
for
the
unsaturated
aliphatic
nitriles
(
e.
g.,
acrylonitrile),
but
not
the
saturated
aliphatic
nitriles
(
e.
g.,
acetonitrile).
This
result,
coupled
with
the
difference
between
the
spectrum
of
dysmorphogenesis
observed
for
the
unsaturated
nitriles
and
for
embryos
treated
with
sodium
cyanide,
led
the
authors
to
suggest
that
microsomal
metabolism
of
unsaturated
nitriles
generates
toxic
metabolites
in
addition
to
cyanide.
In
further
support
of
a
mechanism
distinct
from
cyanide
release
for
the
in
vitro
developmental
toxicity
of
the
tested
compounds,
their
relative
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
9
Final
Draft
potency
in
the
whole­
embryo
culture
assay
did
not
directly
correspond
with
increasing
cyanide
release
kinetics
as
determined
from
in
vitro
metabolism
studies.

No
haloacetontriles
were
included
in
the
study
by
Saillenfait
and
Sabate
(
2000),
but
based
on
their
shared
ability
to
deplete
cellular
GSH
(
as
described
in
the
rest
of
this
paragraph),
HANs
may
act
chemically
more
like
unsaturated
nitriles
than
saturated
ones,
and
thus
may
induce
developmental
toxicity
through
mechanisms
other
than
cyanide
release.
Ahmed
et
al.
(
1982)

reported
that
unsaturated
but
not
saturated
aliphatic
nitriles
decrease
liver
glutathione
levels.

Clinical
signs
of
toxicity
were
different
for
unsaturated
nitriles
than
for
potassium
cyanide,
while
animals
administered
saturated
nitriles
and
potassium
cyanide
showed
a
similar
spectrum
of
symptoms.
Taken
together,
the
results
of
Saillenfait
and
Sabate
(
2000)
and
Ahmed
et
al.
(
1982)

suggest
that
reactive
metabolites
in
addition
to
cyanide
might
be
important
for
the
induction
of
developmental
toxicity.
In
further
support
for
a
common
toxic
mechanism
for
unsaturated
nitriles
and
HANs,
Smith
et
al.
(
1989)
noted
that
a
similar
spectrum
of
soft­
tissue
malformations
is
observed
for
TCAN,
DCAN,
and
the
unsaturated
nitrile,
acrylonitrile.
However,
later
evidence
presented
by
Christ
et
al.
(
1996),
demonstrated
that
the
use
of
tricaprylin
in
these
earlier
studies
might
have
caused
a
shift
in
the
type
of
soft­
tissue
malformations,
placing
the
earlier
conclusion
in
doubt.

Some
results,
however,
do
suggest
an
involvement
of
cyanide
in
the
developmental
toxicity
of
HANs.
In
contrast
to
their
whole­
embryo
assay
results,
Saillenfait
and
Sabate
(
2000)

reported
similarities
in
the
spectrum
of
developmental
effects
induced
by
all
of
the
nitriles
when
tested
in
vivo.
In
this
experiment,
CD­
1
mice
dams
received
a
single
dose
of
aliphatic
nitrile
on
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
10
Final
Draft
gestation
day
10,
and
evaluation
of
embryos
for
defects
was
conducted
on
gestation
day
12.
The
dysmorphogenic
effects
characteristic
of
sodium
cyanide,
including
misdirected
allantois
(
a
tubular
structure
of
the
embryonic
hindgut),
trunk,
or
caudal
extremity
were
induced
by
saturated
as
well
as
unsaturated
aliphatic
nitriles.
The
concordance
in
these
results
in
vivo
is
consistent
with
cyanide
being
an
active
moiety
for
both
saturated
and
unsaturated
aliphatic
nitriles.

Moudgal
et
al.
(
2000)
evaluated
relationships
between
the
structure
of
244
disinfectant
byproducts,
including
21
nitriles,
and
potential
developmental
toxicity
using
a
rat
oral
developmental
toxicity
submodel
of
TOPKAT
®
,
a
quantitative
structure
toxicity
relationship
(
QSTR)
prediction
tool.
Based
on
individual
structural
descriptors,
model
probabilities
were
used
to
derive
qualitative
estimates
as
follows:
0.0
to
0.3
negative,
0.3
to
0.7
indeterminate,
0.7
to
1.0
positive.
The
probability
estimate
is
independent
of
the
potency
or
severity
of
developmental
effects
that
could
be
induced,
and
would
be
interpreted
as
the
likelihood
that
the
chemical
can
cause
developmental
toxicity
in
rats
following
oral
dosing.
As
a
group,
the
nitrile
disinfectant
byproducts
were
characterized
as
having
a
high
probability
of
developmental
toxicity.
Of
the
21
individual
nitrile
compounds,
13
were
positive,
5
were
negative,
and
for
3
the
model
did
not
make
a
prediction.
Only
three
of
the
four
specific
HANs
under
consideration
in
this
document
were
assessed.
DBAN
and
TCAN
were
predicted
as
positive
and
BCAN
was
predicted
as
negative.

TCAN
was
one
of
the
chemicals
in
the
model
training
set.
Four
other
HANs
were
predicted
as
positive.
A
cursory
evaluation
of
these
data
do
not
reveal
a
clear
reason
for
the
negative
prediction
for
BCAN,
since
positive
results
were
obtained
for
bromoacetonitrile,

bromodichloroacetonitrile,
and
dibromochloroacetonitrile.
The
effects
of
individual
structural
moieties
were
also
examined
by
the
software
for
various
structural
classes
of
compounds.
The
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
11
Final
Draft
nitrile
moiety,
chlorine
atom,
and
bromine
atom
were
identified
as
contributing
most
significantly
to
the
developmental
toxicity
predictions
for
the
group
of
20
nitriles.
The
importance
of
the
nitrile
group
in
the
developmental
toxicity
predictions
is
consistent
with
cyanide
as
a
causal
factor
in
the
developmental
toxicity
of
these
compounds
in
vivo,
since
this
functional
group
gives
rise
to
cyanide
during
HAN
metabolism.

In
summary,
mechanisms
of
toxicity
for
noncarcinogenic
effects
(
decreased
body
weight,

gastric
and
liver
toxicity,
and
developmental
effects)
have
been
hypothesized
to
be
related
to
direct
interaction
with
cellular
proteins
(
disrupting
critical
enzyme
functions),
depletion
of
cellular
antioxidant
defenses
(
i.
e,
depletion
of
GSH
and
inhibition
of
GST),
or
effects
secondary
to
cyanide
formation.
The
data
are
not
adequate
to
rule
out
any
of
these
possibilities,
and
therefore,

any
or
all
of
these
mechanisms
could
be
involved.

B.
Mechanism
of
Carcinogenesis
The
carcinogenic
potential
of
the
HANs
is
unknown.
No
epidemiological
studies
have
evaluated
directly
the
carcinogenic
potential
of
HANs
in
humans.
Rather,
studies
have
evaluated
the
carcinogenic
potential
of
chlorinated
versus
unchlorinated
drinking
water
or
the
presence
of
trihalomethanes
as
a
marker
of
chlorination
by­
products
(
IARC,
1999;
Mills
et
al.,
1998).
Many
of
these
studies
have
shown
an
association
between
chronic
exposure
to
chlorinated
water
and
increased
risks
of
bladder,
rectal,
or
colon
cancers
(
Mills
et
al.,
1998;
WHO,
2000).
No
standard
cancer
bioassays
of
HANs
have
been
done
in
animals.
Limited
short­
term
exposure
data
from
the
mouse
skin
assay
(
Bull
et
al.,
1985)
and
the
mouse
lung
assay
(
Bull
and
Robinson,
1985)
indicate
that
BCAN,
DBAN,
and
TCAN
may
be
tumorigenic,
although
DBAN,
DCAN,
and
TCAN
were
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
12
Final
Draft
reported
to
be
negative
in
the
rat
liver
GGT­
foci
assay
(
Herren­
Freund
and
Pereira,
1986).
In
a
qualitative
review
of
structure
activity
relationships,
Bull
and
Robinson
(
1985)
discussed
potential
structural
relationships
among
the
results
for
the
cancer
screening
assays
and
genotoxicity
results
for
various
HANs.
They
commented
that
there
is
no
apparent
consistent
pattern
in
the
potency
of
the
individual
compounds
across
the
various
assays.

Quantitative
analysis
of
structural
relationships
of
HANs
with
carcinogenic
outcomes
have
also
been
investigated.
Moudgal
et
al.
(
2000)
evaluated
relationships
between
the
structure
of
244
disinfectant
byproducts,
including
21
HANs,
and
potential
carcinogenicity
using
mouse
and
rat
oral
submodels
of
TOPKAT
®
,
a
quantitative
structure
toxicity
relationship
(
QSTR)
prediction
tool.
As
a
group,
the
nitrile
disinfectant
byproducts
were
characterized
as
having
a
low
probability
of
carcinogenicity.
However,
for
the
subset
of
six
HANs
in
the
total
group
of
20
nitrile
compounds
for
which
individual
data
were
presented
(
carcinogenicity
predictions
for
TCAN
were
not
included),
four
were
predicted
as
positive
in
at
least
one
sex
in
mice
or
rats.
The
results
were
largely
mixed
across
species
or
sex
for
each
test
compound.
For
example,
BCAN
was
predicted
as
positive
in
male
and
female
mice,
but
negative
in
both
sexes
of
rats.
DBAN
was
negative
in
both
sexes
of
mice,
and
female
rats,
but
indeterminate
in
male
rats.
These
QSTR
results
are
consistent
with
the
screening
bioassays
in
indicating
at
least
limited
carcinogenic
potential
of
the
HANs.

The
HANs
or
their
metabolites
are
reactive
compounds
that
can
bind
macromolecules
including
DNA
(
Daniel
et
al.,
1986;
Lin
et
al.,
1992).
Nouraldeen
and
Ahmed
(
1996)

demonstrated
in
vitro
that
the
degree
of
DNA
binding
was
much
lower
for
DCAN
and
TCAN
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
13
Final
Draft
than
it
was
for
bromoacetonitrile
or
MCAN.
Based
on
changes
in
fluorescence
as
a
measure
of
adduct
formation,
relative
DNA
adduct
formation
as
compared
to
bromoacetonitrile
was
8.6%,

1.0%,
and
0.2%
for
MCAN,
DCAN,
and
TCAN
respectively.
The
chemical
nature
of
the
adducts
that
were
formed
was
not
identified
for
each
compound.
However,
a
7­

(
cyanomethyl)
guanine
adduct
was
the
single
major
adduct
identified
from
the
reaction
of
bromoacetonitrile
with
calf­
thymus
DNA.
Since
the
more
fully
halogenated
acetonitriles
BCAN,

DBAN,
DCAN,
and
TCAN,
may
be
metabolized
to
monohaloacetonitriles
in
vivo
(
Pereira
et
al.,

1984),
they
might
induce
the
formation
of
similar
adducts.
On
the
other
hand,
the
relative
DNA
reactivity
observed
in
vitro
may
not
directly
translate
to
mutagenic
or
carcinogenic
potential
in
vivo,
since
the
metabolism
of
the
compound
and
the
mutagenicity
of
the
adducts
formed
may
differ
for
each
HAN.

Glutathione
conjugation
may
be
an
important
cellular
protection
against
these
reactive
HANs
or
their
metabolites
(
Lin
and
Guion,
1989;
Ahmed
et
al.,
1991).
The
genotoxicity
of
each
of
the
HAN
compounds
BCAN,
DBAN,
DCAN,
and
TCAN
has
been
evaluated
in
at
least
one
assay,
and
in
most
cases
a
variety
of
different
assays.
Although
some
of
the
data
have
provided
contradictory
results,
all
of
the
tested
compounds
appear
to
have
some
capacity
to
induce
genotoxic
effects.
For
example,
BCAN,
DCAN,
and
TCAN
(
but
not
DBAN)
have
been
found
to
generate
a
positive
result
in
at
least
one
reported
Salmonella/
microsome
assay.
In
addition,
while
not
uniformly
consistent,
a
variety
of
other
assays,
including
those
for
chromosome
effects,
have
yielded
positive
results
for
some
of
the
HANs.
Overall,
these
data
suggest
that
HANs
induce
genotoxicity
through
direct
interactions
with
DNA.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
14
Final
Draft
C.
Interactions
and
Susceptibilities
Potential
Interactions
No
studies
on
interactions
of
HANs
with
other
classes
of
compounds
were
identified,

except
as
noted
above
for
solvent
vehicle
effects.

Childhood
Susceptibility
As
discussed
above,
developmental
toxicity
has
been
associated
with
exposure
to
BCAN,

DBAN,
DCAN,
and
TCAN,
although
the
findings
for
all
these
compounds,
except
TCAN,
are
confounded
by
the
developmental
toxicity
of
the
vehicle.
The
developmental
toxicity
of
the
HANs
is
supported
by
the
reports
of
Roth
et
al.
(
1990)
and
Gordon
et
al.
(
1991)
in
published
abstracts
that
radioactivity
was
detected
in
embryos
of
dams
given
[
14C]
TCAN,
suggesting
that
fetuses
may
be
targets
for
HAN
toxicity.
In
addition,
Abdel­
Aziz
et
al.
(
1993)
found
that
fetuses
may
be
more
susceptible
than
adults
to
direct
DNA
damage
induced
by
haloacetontriles.

Although
the
relationship
between
metabolism
of
HANs
and
the
onset
of
toxicity
has
not
been
thoroughly
determined
(
as
discussed
above),
acute
cyanide
intoxication
from
accidental
pediatric
acetonitrile
exposures
(
Caravati
et
al.,
1988;
Geller
et
al.,
1991;
Kurt
et
al.,
1991)
indicates
that
HAN
metabolism
to
cyanide
occurs
in
children.

There
are
limited
data
available
for
assessing
directly
the
susceptibility
of
fetuses
and
children
to
HANs.
No
systemic
toxicity
studies
have
been
identified
that
evaluated
age­
related
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
15
Final
Draft
differences
in
sensitivity
and
no
multigeneration
reproductive
studies
have
been
reported.
In
addition,
age­
related
differences
in
the
metabolism
of
the
HANs
could
not
be
investigated,

because
the
enzyme(
s)
responsible
for
HAN
metabolism
are
not
known.
Pereira
et
al.
(
1984)

hypothesized
that
mixed
function
oxidases
are
involved,
but
the
isozyme
involved
has
not
been
identified,
and
the
age­
dependent
expression
differs
among
the
different
cytochrome
P450
isozymes.
Although
the
developmental
toxicity
of
the
HANs
has
been
evaluated,
most
of
the
available
literature
cannot
be
used
to
determine
relative
maternal
and
developmental
toxicity
of
these
compounds,
due
to
potential
interactions
between
the
test
compound
and
the
dosing
vehicle
(
tricaprylin).
Excluding
these
studies
from
this
evaluation
leaves
only
limited
data.
No
developmental
toxicity
was
observed
following
administration
of
up
to
10.9
mg/
kg/
day
DBAN
in
drinking
water
(
R.
O.
W.
Sciences,
1997),
but
this
was
only
a
screening
study
that
did
not
include
evaluation
of
pups
for
malformations.
The
only
maternal
effect
at
this
dose
was
a
decrease
in
drinking
water
consumption.
In
a
study
using
corn
oil
as
the
solvent
vehicle
for
TCAN,
the
maternal
NOAEL
was
15
mg/
kg/
day
and
the
NOAEL
for
developmental
toxicity
was
35
mg/
kg/
day
(
Christ
et
al.,
1996).
Thus,
these
two
studies
do
not
provide
evidence
that
fetuses
are
more
susceptible
than
adults,
although
the
data
are
too
limited
to
make
a
definitive
conclusion.

Other
potential
susceptibilities
Although
a
role
of
oxidative
metabolism
and
glutathione
conjugation
have
been
hypothesized
to
be
involved
in
HAN
metabolism,
the
identity
of
enzymes
involved
in
these
pathways
has
not
been
determined.
Therefore,
the
potential
role
of
interindividual
differences
in
metabolism
due
to
genetic
polymorphism,
age,
or
other
factors
cannot
be
determined.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
16
Final
Draft
D.
Summary
The
HANs
induce
general
systemic
toxicity.
Decreased
body
weight
and
a
variety
of
organ
weight
changes
occur
following
oral
dosing,
and
the
testes
(
NTP,
2002)
and
liver
might
be
particularly
sensitive
(
Hayes
et
al.,
1986),
although
the
reported
effects
in
these
organs
in
available
studies
are
fairly
limited.
The
HANs
also
induce
developmental
effects
(
Smith
et
al.,

1986;
Smith
et
al.,
1987;
Smith
et
al.,
1988;
Smith
et
al.,
1989;
Christ
et
al.,
1995;
Christ
et
al.,

1996).
The
mechanism(
s)
of
toxicity
are
not
known,
but
several
possibilities
have
been
described.

HANs
may
act
through
direct
interactions
with
cellular
macromolecules
such
as
DNA
(
Daniel
et
al.,
1986;
Lin
et
al.,
1992;
Nouraldeen
and
Ahmed,
1996).
HAN
toxicity
might
be
secondary
to
GSH
depletion
(
Ahmed
et
al.,
1991)
or
oxidative
stress
(
Ahmed
et
al.,
1999;
Mohamadin
and
Abdel­
Naim,
1999).
Formation
of
cyanide
from
HAN
might
be
another
important
mechanism
of
toxicity,
although
important
systemic
effects
that
are
sensitive
indicators
of
cyanide
toxicity
have
not
been
fully
examined.

The
role
of
cyanide
in
the
developmental
toxicity
of
HANs
has
received
much
attention.

Some
studies
suggest
that
metabolites
other
than
cyanide
play
a
critical
role
(
Smith
et
al.,
1986),

and
implicated
glutathione
depletion
as
an
important
factor
(
Christ
et
al.,
1995;
Abdel­
Aziz
et
al.,

1993).
Although
some
indirect
data
supports
a
role
of
cyanide
(
Moudgal
et
al.,
2000;
Saillenfait
and
Sabate,
2000),
evaluation
of
the
available
developmental
toxicity
studies
of
cyanide
itself
do
not
support
this
hypothesis
(
U.
S.
EPA,
2002c).

The
ability
of
the
HANs
to
bind
to
cellular
macromolecules
(
Daniel
et
al.,
1986;
Lin
et
al.,

1992;
Nouraldeen
and
Ahmed,
1996),
as
well
as
generally
positive
results
in
genotoxicity
assays,
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
VII­
17
Final
Draft
supports
direct
DNA
damage
as
the
mode
of
action
for
the
tumorigenicity
observed
in
cancer
screening
studies
(
Bull
et
al.,
1985;
Bull
and
Robinson,
1985).
However,
the
carcinogenic
potential
of
the
HANs
is
unknown,
since
epidemiology
studies
are
not
available
and
standard
cancer
animal
bioassays
of
HANs
have
not
been
conducted.

Identification
of
potential
susceptible
subpopulations
is
hampered
by
the
incomplete
characterization
of
HAN
metabolism
or
identification
of
the
toxic
moiety.
Although
a
metabolic
pathway
for
the
HANs
has
been
proposed
(
Pereira
et
al.,
1984),
the
enzymes
important
for
catalyzing
HAN
metabolism
are
unknown.
In
addition,
no
studies
on
age­
dependent
differences
in
metabolism
or
toxicity
were
identified,
although
one
study
demonstrated
that
HANs
may
bind
more
greatly
to
fetal
DNA
than
to
DNA
in
maternal
tissues
(
Abdel­
Aziz
et
al.,
1993).
Analysis
of
the
developmental
toxicity
studies
for
TCAN
revealed
a
lower
maternal
than
developmental
NOAEL,
which
does
not
suggest
that
fetuses
are
more
susceptible
than
adults.