Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
I­
1
Final
Draft
Chapter
I.
Executive
Summary
A.
Introduction
Haloacetonitriles
(
HANs)
are
derivatives
of
acetonitrile
(
CH
3
CN),
in
which
one
to
three
halogen
atoms
are
substituted
for
hydrogen.
The
four
halogenated
acetonitriles
selected
for
consideration
in
this
document
are
bromochloroacetonitrile
(
BCAN),
dibromoacetonitrile
(
DBAN),
dichloroacetonitrile
(
DCAN),
and
trichloroacetonitrile
(
TCAN).
These
HANs
were
selected
for
inclusion
in
this
document
in
consideration
of
the
prevalence
of
individual
HANs
in
drinking
water,
and
the
availability
of
toxicity
data.
This
document
includes
an
evaluation
of
literature
on
the
HANs
resulting
from
a
full
literature
search
for
toxicity
data
conducted
in
December
1999,
and
exposure
data
in
May
2002.
Key
newer
studies
identified
after
the
literature
search
date
have
been
included
as
available
at
the
time
of
document
preparation.

B.
Human
Exposure
The
Information
Collection
Rule
(
ICR)
database
(
U.
S.
EPA,
2002a)
contains
extensive
information
on
concentrations
of
BCAN,
DBAN,
DCAN,
and
TCAN
in
drinking­
water
systems,

and
on
how
those
concentrations
vary
with
input­
water
characteristics
and
treatment
methods.

The
database
contains
information
from
six
quarterly
samples
from
7/
97
to
12/
98,
from
approximately
300
large
systems
covering
approximately
500
plants.
The
mean
concentrations
of
BCAN
were
0.73
and
1.14

g/
L
in
groundwater
and
surface
water,
respectively.
The
mean
concentrations
of
DBAN
were
0.82
and
0.75

g/
L
in
groundwater
and
surface
water,

respectively.
The
mean
concentrations
of
DCAN
were
0.87
and
2.21

g/
L
in
groundwater
and
surface
water,
respectively.
The
mean
concentrations
of
TCAN
were
0.14
and
0.03

g/
L
in
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
I­
2
Final
Draft
groundwater
and
surface
water,
respectively.
The
median
concentrations
of
BCAN,
DBAN,

DCAN,
and
TCAN
were
less
than
their
means
in
groundwater
and
surface
water.

HANs
are
produced
during
water
chlorination
or
chloramination
from
naturally
occurring
substances,
including
algae,
humic
acid,
fulvic
acid,
and
proteinaceous
material.
Reckhow
et
al.

(
1990)
found
that
disinfection
of
water
containing
humic
acids
resulted
in
higher
concentrations
of
HANs
than
disinfection
of
water
containing
the
corresponding
fulvic
acids.

The
disinfection
process
producing
the
highest
concentration
of
HANs
was
chlorination.

Chloramine
produced
lower
levels
of
HANs.
Most
investigators
(
Boorman
et
al.,
1999;

Richardson
1998;
Lykins
et
al.,
1994;
Jacangelo
et
al.,
1989)
found
that
the
formation
of
HANs
when
ozonation
was
followed
by
chlorine
or
chloramine
was
less
than
when
chlorine
or
chloramine
was
the
sole
disinfectant.
Interestingly,
Miltner
et
al.
(
1990)
reported
that
the
formation
of
DBAN
in
simulated
distribution
water
was
higher
(
p
=
0.05)
when
ozonation
was
combined
with
chlorination
or
with
chloramination
than
when
chlorination
was
used
alone.
In
addition,
Miltner
et
al.
(
1990)
found
that
ozonation
had
no
statistically
significant
effect
on
the
formation
of
BCAN,
DCAN,
or
TCAN.
Richardson
(
1998)
found
that
BCAN,
DCAN,
and
TCAN
were
not
produced
in
measurable
quantities
by
ozonation
or
chlorine
dioxide.
However,

DBAN
was
formed
by
ozone
in
the
presence
of
elevated
bromide,
but
not
by
chlorine
dioxide
disinfection.

Ambient
bromide
levels
appear
to
influence,
to
some
degree,
the
speciation
of
HANs.

DCAN
is
by
far
the
most
predominant
HAN
detected
in
drinking
water
from
sources
with
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
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Final
Draft
bromide
levels
of
20
µ
g/
L
or
less.
In
treated
water
from
sources
with
higher
bromide
levels
(
50
 
80

g/
L),
BCAN
was
the
second
most
prevalent
compound
(
WHO,
2000).
Richardson
(
1998)
found
that
when
bromide
was
present
in
the
source
water,
DBAN
concentrations
were
greater
than
those
of
chloroform
or
dichloroacetic
acid,
which
normally
predominate.

In
general,
increasing
temperature
and/
or
decreasing
pH
has
been
associated
with
increasing
concentrations
of
HANs
(
AWWARF,
1991;
Siddiqui
&
Amy,
1993).
Although
HANs
form
rapidly,
they
decay
in
the
distribution
system
as
a
result
of
hydrolysis.
HANs
hydrolyzed
at
pH
levels
>
9.0
and
continued
to
degrade
in
the
distribution
system
(
Arora
et
al.,
1997).
The
relative
stability
of
individual
HANs
appears
to
be
dependent
on
the
specific
source
water
(
AWWARF,
1991).

In
general,
there
were
no
clear
trends
of
the
concentrations
of
HANs
with
season.

However,
among
35
water
treatment
facilities
investigated,
Krasner
et
al.
(
1989)
found
that
at
the
facility
with
the
highest
bromide
level
(~
3
mg/
L
bromide),
there
was
a
shift
in
the
distribution
of
HANs
from
chlorinated
HANs
to
brominated
HANs.

TCAN
has
been
used
as
an
insecticide
(
Budavari
et
al.,
1989).
No
data
were
located
on
exposure
to
BCAN,
DBAN,
DCAN,
and
TCAN
in
food,
air,
or
via
dermal
exposure
when
showering
or
swimming.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
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Final
Draft
C.
Toxicokinetics
Limited
data
are
available
on
the
toxicokinetics
of
the
HANs,
with
a
comprehensive
toxicokinetic
study
for
oral
dosing
available
only
for
DCAN.
However,
the
existing
toxicokinetic
data
suggest
that
HANs
can
be
rapidly
and
nearly
completely
absorbed
following
oral
dosing
(
Roby
et
al.,
1986;
Roth
et
al.,
1990).
Systemic
toxicity
data
suggest
that
HANs
are
absorbed
by
the
dermal
route.
Once
absorbed,
HANs
appear
to
be
widely
distributed.
The
two
compounds
tested,
DCAN
(
Roby
et
al.,
1986)
and
TCAN
(
Lin
et
al.,
1992),
were
widely
distributed
following
oral
dosing,
with
no
clear
preferences
in
tissue
distribution
apparent
based
on
the
limited
data.
No
data
were
available
on
tissue­
dependent
metabolism,
but
an
overall
metabolic
scheme
for
HANs
involving
an
initial
oxidative
dehalogenation
step
has
been
proposed
based
on
the
ability
of
these
compounds
to
form
cyanide
and
metabolism
studies
for
other
nitriles
(
Pereira
et
al.,
1984).
Proposed
intermediate
metabolites
have
not
been
measured
directly,
and
the
identity
of
enzymes
responsible
for
steps
in
the
pathway
have
not
been
identified.
Conjugation
with
glutathione
(
GSH),
at
least
at
high
doses,
might
be
a
second
important
route
of
metabolism
for
HANs
(
Ahmed
et
al.,
1989;
Lin
and
Guion,
1989;
Ahmed
et
al.,
1991;
NTP,
2002).

Excretion
of
HANs
is
nearly
complete
over
a
period
of
days,
largely
in
urine
and
in
exhaled
air.

The
rate
of
excretion
may
differ
across
species,
since
mice
excrete
DCAN
more
rapidly
than
rats
(
Roby
et
al.,
1986).
Differences
in
urinary
excretion
of
thiocyanate
for
different
HANs
was
observed
by
Pereira
et
al.
(
1984),
with
TCAN
being
excreted
as
thiocyanate
to
a
lesser
degree
than
the
other
HANs.
The
results
of
Roby
et
al.
(
1986)
that
showed
relatively
rapid
excretion
of
DCAN­
associated
radioactivity
suggests
limited
potential
for
the
bioaccumulation
of
the
HANs.

However,
no
studies
were
located
that
provided
data
on
long­
term
accumulation
and
retention
of
any
of
the
HANs.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
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HECD
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5
Final
Draft
D.
Health
Effects
In
Animals
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.

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
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
I­
6
Final
Draft
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.

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
Drinking
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Document
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Final
Draft
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
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.
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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.

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
gamma­
glutamyltranspeptidase
(
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
EPA/
OW/
OST/
HECD
I­
9
Final
Draft
E.
Health
Effects
in
Humans
Human
epidemiology
data
on
the
toxicity
of
the
HANs
are
lacking.
Most
of
the
human
health
data
for
HANs
are
as
components
of
complex
mixtures
of
water
disinfection
byproducts.

These
complex
mixtures
of
disinfection
byproducts
have
been
associated
with
increased
potential
for
adverse
effects
on
reproduction
(
reviewed
by
Nieuwenhuijsen
et
al.,
2000).
Although
most
studies
of
human
health
effects
following
exposure
to
water
disinfectant
byproducts
have
used
total
trihalomethanes
as
the
exposure
metric,
Klotz
and
Pyrch
(
1999),
conducted
a
case­
control
study
on
the
relationship
between
neural
tube
defects
and
drinking
water
exposure
to
trihalomethanes,
HANs,
and
haloacetic
acids.
The
specific
compounds
that
were
measured
as
part
of
the
total
HAN
exposure
estimate
were
not
identified.
Based
on
the
results
of
the
study,

the
authors
concluded
that
the
HANs
did
not
exhibit
a
clear
association
with
neural
tube
defects.

No
epidemiological
studies
have
evaluated
directly
the
carcinogenic
potential
of
HANs
in
humans.
Rather,
studies
have
evaluated
the
carcinogenic
potential
of
chlorinated
versus
unchlorinated
drinking
water
or
the
presence
of
trihalomethanes
as
a
marker
of
chlorination
byproducts
(
IARC,
1999;
Mills
et
al.,
1998).
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).

F.
Mechanisms
of
Toxicity
and
Sensitive
Subpopulations
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
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
I­
10
Final
Draft
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,

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.
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
I­
11
Final
Draft
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.

G.
Derivation
of
the
Health
Advisories
Health
Advisory
values
(
HA)
for
BCAN,
DBAN,
DCAN,
and
TCAN
are
summarized
in
Table
I­
1
and
the
derivation
of
these
values
is
shown
in
Chapter
VIII.
Based
on
the
clear
limitations
in
the
database
and
gaps
in
understanding
of
the
mechanisms
of
toxicity
for
HANs,
the
derived
RfD
and
HA
values
are
best
characterized
as
low
in
confidence.
For
BCAN,
no
suitable
studies
were
identified
for
derivation
of
any
HAs.
For
DBAN,
no
suitable
studies
were
located
for
derivation
of
a
One­
day
HA.
A
NOAEL
of
12
mg/
kg/
day
for
decreased
body
weight,

decreased
testes
weight
and
testes
atrophy
in
male
F344
rats
in
a
14­
day
drinking
water
study
(
NTP,
2002)
was
used
to
derive
a
Ten­
day
HA
value
of
1
mg/
L
for
a
10­
kg
child.
This
Ten­
day
HA
was
used
as
a
conservative
value
for
the
One­
day
HA.
A
NOAEL
of
11.3
mg/
kg/
day
was
also
identified
in
a
parallel
90­
day
drinking
water
study
in
male
F344
rats
(
NTP,
2002).
The
NOAEL
value
was
used
to
calculate
the
Longer­
term
HA
value
of
0.4
mg/
L
for
a
10­
kg
child
and
1
mg/
L
for
a
70­
kg
adult.
No
chronic
study
of
DBAN
toxicity
was
located,
so
the
subchronic
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
I­
12
Final
Draft
(
13­
week)
NOAEL
was
employed
with
an
extra
uncertainty
factor
to
account
for
extrapolation
from
subchronic
to
chronic
exposure
to
calculate
a
Life­
time
HA
value
of
0.08
mg/
L
(
80
µ
g/
L).

Table
I­
1.
Summary
of
Health
Advisory
Values
for
Drinking
Water
(
a)

Longer­
Term
HA
Chemical
One­
Day
HA
Ten­
Day
HA
Child
Adult
Life­
Time
HA
BCAN
 
(
b)
 
­
­
 
 
DBAN
1
1
0.4
1
0.08
DCAN
0.4
(
0.5c)
0.4
(
0.5c)
0.03
(
0.04c)
0.09
(
0.1c)
0.02
(
0.009c)

TCAN
2
(
2c)
2
(
2
c
)
­­
­­
­­

a
mg/
L
b
No
value
calculated
due
to
lack
of
suitable
toxicological
data.
c
Value
calculated
from
the
study
BMDL.

For
DCAN,
no
suitable
studies
were
located
for
derivation
of
a
One­
day
HA.
A
LOAEL
of
12
mg/
kg/
day
for
increased
relative
liver
weight
in
male
rats
greater
than
10%
in
a
14­
day
gavage
study
(
Hayes
et
al.,
1986)
was
used
to
derive
a
Ten­
day
HA
value
of
0.4
mg/
L
for
a
10­

kg
child.
Further
analysis
of
these
data
yielded
a
BMDL
of
5
mg/
kg/
day
as
the
critical
effect
level
for
this
same
effect.
When
derived
on
the
basis
of
the
BMDL,
the
Ten­
day
HA
value
is
0.5
mg/
L.

The
Ten­
day
HA
was
used
as
a
conservative
value
for
the
One­
day
HA.
A
LOAEL
of
8
mg/
kg/
day
was
identified
for
increased
relative
liver
weight,
supported
by
clinical
chemistry
findings
at
higher
doses
in
a
90­
day
study
in
male
and
female
rats
(
Hayes
et
al.,
1986).
Further
analysis
of
these
data
yielded
a
BMDL
of
4
mg/
kg/
day
as
the
critical
effect
level.
The
LOAEL
and
BMDL
were
used
to
calculate
the
Longer­
term
HA
value.
When
derived
on
the
basis
of
the
LOAEL,
the
Longer­
term
HA
value
was
0.03
mg/
L
for
a
10­
kg
child
and
was
0.09
mg/
L
for
a
70­
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
I­
13
Final
Draft
kg
adult.
When
derived
on
the
basis
of
the
BMDL,
the
Longer­
term
HA
value
was
0.04
mg/
L
for
a
10­
kg
child
and
0.1
mg/
L
for
the
70­
kg
adult.
No
chronic
study
of
DCAN
toxicity
was
located,

so
the
subchronic
(
90­
day)
LOAEL
and
BMDL
were
employed
with
an
extra
uncertainty
factor
to
calculate
the
Life­
time
HA
values
of
0.02
mg/
L
(
based
on
the
LOAEL)
and
0.009
mg/
L
(
based
on
the
BMDL).

For
TCAN,
no
suitable
studies
were
located
for
derivation
of
a
One­
day
HA.
A
NOAEL
of
15
mg/
kg/
day
for
absence
of
a
decrease
in
maternal
body
weight
gain
in
pregnant
rats
was
used
to
derive
a
Ten­
day
HA
value
of
2
mg/
L
for
a
10­
kg
child.
Further
analysis
of
these
data
yielded
a
BMDL
of
17
mg/
kg/
day
as
the
critical
effect
level.
When
derived
on
the
basis
of
the
BMDL,

the
Ten­
day
HA
value
is
2
mg/
L.
The
Ten­
day
HA
was
used
as
a
conservative
value
for
the
Oneday
HA.
No
suitable
subchronic
or
chronic
toxicity
data
were
located
for
derivation
of
Longerterm
or
Life­
time
HAs.

Due
to
lack
of
adequate
dose­
response
information,
calculations
of
carcinogenic
risk
have
not
been
performed
for
any
of
the
haloacetonitriles.
The
limited
short­
term
data
from
the
mouse
lung
and
skin
assays,
as
well
as
QSTR
analyses,
indicate
that
BCAN,
DBAN,
and
TCAN
may
have
some
carcinogenic
potential
in
animals.
In
addition,
data
suggest
that
haloacetonitriles
may
induce
genotoxicity
through
direct
interactions
with
DNA.
Although
the
available
data
provide
at
least
limited
indications
of
potential
carcinogenicity
of
HANs,
these
data
are
not
adequate
to
demonstrate
carcinogenicity
in
animals.
Following
EPA's
1986
Guidelines
for
Cancer
Risk
Assessment
(
U.
S.
EPA,
1986),
BCAN,
DBAN,
DCAN,
and
TCAN
are
appropriately
classified
as
Group
D
­
Not
Classifiable
as
to
Human
Carcinogenicity.
This
classification
is
appropriate
Drinking
Water
Criteria
Document
for
Haloacetonitriles
EPA/
OW/
OST/
HECD
I­
14
Final
Draft
when
there
is
inadequate
evidence
of
carcinogenicity
in
humans
or
animals.
Following
EPA's
Draft
1999
Guidelines
for
Cancer
Risk
Assessment
(
U.
S.
EPA,
1999),
the
data
for
the
HANs
can
best
be
described
as
Data
Are
Inadequate
for
an
Assessment
of
Human
Carcinogenic
Potential.
