Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
Final
draft
XIII­
1
Chapter
VIII.
Quantification
of
Toxicological
Effects
The
quantification
of
toxicological
effects
of
a
chemical
consists
of
separate
assessments
of
noncarcinogenic
and
carcinogenic
health
effects.
Unless
otherwise
specified,
chemicals
which
do
not
produce
carcinogenic
effects
are
believed
to
have
a
threshold
dose
below
which
no
adverse,
noncarcinogenic
health
effects
occur,
while
carcinogens
are
assumed
to
act
without
a
threshold.

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

RfD
=
NOAEL
(
LOAEL)
UF
×
MF
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
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EPA/
OW/
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HECD
Final
draft
XIII­
2
where:

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

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

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

The
individual
components
of
the
uncertainty
are
as
follows:

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

UF
A
An
additional
factor
of
1,
3,
or
10
is
used
when
extrapolating
from
valid
results
of
long­
term
studies
on
experimental
animals
when
results
of
studies
of
human
exposure
are
not
available
or
are
inadequate.
This
factor
is
intended
to
account
for
the
uncertainty
involved
in
extrapolating
from
animal
data
to
humans
(
interspecies
variation).
Drinking
Water
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Document
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Cyanogen
Chloride
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EPA/
OW/
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HECD
Final
draft
XIII­
3
UF
S
An
additional
factor
of
1,
3,
or
10
is
used
when
extrapolating
from
lessthan
chronic
results
on
experimental
animals
when
there
are
no
useful
longterm
human
data.
This
factor
is
intended
to
account
for
the
uncertainty
involved
in
extrapolating
from
less­
than­
chronic
NOAELs
to
chronic
NOAELs.

UF
L
An
additional
factor
of
1,
3,
or
10
is
used
when
deriving
an
RfD
from
a
LOAEL,
instead
of
a
NOAEL.
This
factor
is
intended
to
account
for
the
uncertainty
involved
in
extrapolating
from
LOAELs
to
NOAELs.

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

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

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

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

Unless
otherwise
specified,
chemicals
which
do
not
produce
carcinogenic
effects
are
believed
to
have
a
threshold
dose
below
which
no
adverse,
noncarcinogenic
health
effects
occur,
while
carcinogens
are
assumed
to
act
without
a
threshold.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
Final
draft
XIII­
4
A.
1.1.
Drinking
Water
Equivalent
Level
The
drinking
water
equivalent
(
DWEL)
is
calculated
from
the
RfD.
The
DWEL
represents
a
drinking­
water­
specific
lifetime
exposure
at
which
adverse,
noncarcinogenic
health
effects
are
not
anticipated
to
occur.
The
DWEL
assumes
100%
exposure
from
drinking
water.

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

DWEL
=
(
RfD)
×
BW
WI
where:

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

A.
1.2.
Health
Advisory
Values
In
addition
to
the
RfD
and
the
DWEL,
EPA
calculates
Health
Advisory
(
HA)
values
for
noncancer
effects.
HAs
are
determined
for
lifetime
exposures
as
well
as
for
exposures
of
shorter
duration
(
1­
day,
10­
day,
and
longer­
term).
The
shorter
duration
HA
values
are
used
as
informal
guidance
to
municipalities
and
other
organizations
when
emergency
spills
or
contamination
situations
occur.
The
lifetime
HA
becomes
the
MCLG
for
a
chemical
that
is
not
a
carcinogen.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
Final
draft
XIII­
5
The
shorter­
term
HAs
are
calculated
using
an
equation
similar
to
the
ones
for
RfD
and
DWEL;
however,
the
NOAELs
or
LOAELs
are
derived
from
acute
or
subchronic
studies
and
identify
a
sensitive
noncarcinogenic
endpoint
of
toxicity.
The
HAs
are
derived
as
follows:

HA
=
NOAEL
or
LOAEL
×
BW
UF
×
WI
where:

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

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

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

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

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

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

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

Each
of
these
shorter­
term
HA
values
assumes
that
the
total
exposure
to
the
contaminant
comes
from
drinking
water.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
Final
draft
XIII­
6
The
lifetime
HA
is
calculated
from
the
DWEL
and
takes
into
account
exposure
from
sources
other
than
drinking
water.
It
is
calculated
using
the
following
equation:

Lifetime
HA
=
DWEL
×
RSC
where:

DWEL
=
Drinking
water
equivalent
level
RSC
=
Relative
source
contribution.
The
fraction
of
the
total
exposure
allocated
to
drinking
water.

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

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

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

°
Group
C:
Possible
Human
Carcinogen.
Limited
evidence
of
carcinogenicity
in
animals
in
the
absence
of
human
data.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
Final
draft
XIII­
7
°
Group
D:
Not
classified
as
to
Human
Carcinogenicity.
Inadequate
human
and
animal
evidence
of
carcinogenicity
or
for
which
no
data
are
available.

°
Group
E:
Evidence
of
Noncarcinogenicity
for
Humans.
No
evidence
of
carcinogenicity
in
at
least
two
adequate
animal
tests
in
different
species
or
in
both
adequate
epidemiologic
and
animal
studies.

If
toxicological
evidence
leads
to
the
classification
of
the
contaminant
as
a
genotoxic,

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

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

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

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

Excess
cancer­
risk
estimates
may
also
be
calculated
using
other
models
such
as
the
one­
hit,
Drinking
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Cyanogen
Chloride
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OW/
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HECD
VIII­
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Final
draft
Weibull,
logit
and
probit
models.
There
is
little
basis
in
the
current
understanding
of
the
biological
mechanisms
involved
in
cancer
to
suggest
that
any
one
of
these
models
is
able
to
predict
risk
more
accurately
than
any
of
the
others.
Because
each
model
is
based
upon
differing
assumptions,
the
estimates
that
are
derived
for
each
model
can
differ
by
several
orders
of
magnitude.

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

The
quantification
of
toxicological
effects
of
a
chemical
consists
of
separate
assessments
of
noncarcinogenic
and
carcinogenic
health
effects.
Chemicals
that
do
not
produce
carcinogenic
effects
are
believed
to
have
a
threshold
dose
below
which
no
adverse,
noncarcinogenic
health
effects
occur,
while
carcinogens
are
assumed
to
act
without
a
threshold.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
9
Final
draft
B.
Noncarcinogenic
Effects
The
useful
toxicity
data
for
the
cyanogen
chloride
and
its
known
metabolites
(
cyanide,

thiocyanate)
are
provided
in
Tables
VIII­
1
through
VIII­
4.
(
The
useful
toxicity
data
for
the
putative
metabolites
cyanate
and
cyanamide,
and
for
HCl,
which
is
not
appropriate
as
a
surrogate,

are
summarized
in
Appendix
Tables
C­
1,
D­
1,
and
E­
1,
respectively.)
As
noted
in
the
Executive
Summary,
because
of
the
lack
of
adequate
toxicity
for
cyanogen
chloride,
use
of
data
from
known
or
potential
cyanogen
chloride
metabolites
was
considered
as
the
basis
for
the
cyanogen
chloride
Health
Advisories.
Therefore,
this
section
first
provides
the
derivation
of
Health
Advisories
for
cyanide
and
thiocyanate,
the
known
metabolites
of
cyanogen
chloride.
The
Health
Advisories
for
cyanogen
chloride
are
then
derived,
based
on
consideration
of
the
Health
Advisories
derived
for
the
other
chemicals,
as
well
as
consideration
of
the
limited
toxicity
and
toxicokinetics
data
available
for
cyanogen
chloride.

B.
1.
Cyanide
B.
1.1
One­
day
Health
Advisory
for
Cyanide
No
suitable
studies
were
located.
Only
human
poisoning
studies,
in
which
severe
effects
were
observed
(
e.
g.,
Liebowitz
and
Schwarz,
1948;
Saincher
et
al.,
1994),
and
animal
lethality
studies
(
e.
g.,
Ferguson,
1962;
Smyth
et
al.,
1969),
were
identified.
In
the
absence
of
adequate
data,
the
ten­
day
HA
value
is
recommended
as
a
conservative
estimate
of
an
appropriate
one­
day
HA
value.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
10
Final
draft
B.
1.2.
Ten­
day
Health
Advisory
for
Cyanide
Two
suitable
studies
of
appropriate
duration
were
located.
The
first
was
that
of
Kreutler
et
al.
(
1978).
In
this
study,
rats
were
exposed
in
the
diet
for
2
weeks
to
87
mg
CN/
kg­
day,
and
were
provided
diets
that
either
supplied
adequate
protein
or
that
were
deficient
in
protein
and
iodine.
No
effect
was
seen
in
the
group
receiving
adequate
protein
and
iodine,
but
decreased
plasma
TSH
and
increased
thyroid
weight
were
observed
in
the
group
provided
the
diet
deficient
in
protein
and
iodine.
Although
changes
in
TSH
and
thyroid
weight
alone
can
be
considered
adaptive
rather
than
adverse
effects,
this
study
did
not
include
assessment
of
thyroid
histopathology
to
determine
if
thyroid
hyperplasia
was
present.
Thus,
one
can
consider
this
study
as
identifying
a
minimal
LOAEL
of
40
mg
CN/
kg­
day
in
a
sensitive
population.
Benchmark
dose
analysis
of
TSH
levels
resulted
in
a
BMDL
of
2.1
mg/
kg­
day.
Confidence
in
this
BMDL
is
limited,
however,
because
only
a
single
dose
level
and
a
control
were
available,
meaning
that
only
a
linear
model
could
be
fit
to
the
data.
In
addition
to
the
absence
of
information
on
the
shape
of
the
dose­
response
curve,
the
BMDL
is
markedly
lower
than
the
NOAEL,
meaning
that
calculating
the
BMDL
involved
significant
extrapolation
below
the
data,
a
practice
generally
not
needed
for
BMD
modeling.
Thus,
although
this
BMDL
can
be
considered
a
NOAEL
surrogate
because
the
modeled
effect
is
an
adaptive
effect,
overall
confidence
in
the
BMDL
is
limited.

In
the
second
study
(
Palmer
and
Olson,
1979),
no
effect
was
seen
in
rats
administered
8
mg
CN/
kg­
day
in
feed
for
21
days.
Analysis
of
the
feed
by
the
study
authors
led
them
to
conclude
that
<
20%
of
the
administered
dose
of
cyanide
was
bioavailable;
therefore,
the
NOAEL
can
be
estimated
as
1.6
mg
CN/
kg­
day.
In
a
parallel
study
by
the
same
authors,
a
LOAEL
of
14
mg
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
11
Final
draft
CN/
kg­
day
was
identified
in
rats
administered
cyanide
in
drinking
water
for
the
same
duration,

based
on
increased
liver
weight;
no
drinking
water
NOAEL
corresponding
to
this
LOAEL
was
identified.
BMDL
modeling
could
not
be
conducted
for
this
study,
because
no
measure
of
variability
was
provided.

It
is
of
interest
that
the
doses
in
the
two
short­
term
studies
are
higher
than
the
doses
reported
to
cause
death
in
rats
and
mice
when
delivered
in
a
single
dose
(
4­
8
mg/
kg,
Ferguson,

1962;
Smyth
et
al.,
1969).
However,
this
finding
is
consistent
with
the
mechanism
of
action
of
cyanide.
A
bolus
dose
results
in
rapid
absorption
of
cyanide,
which
can
overwhelm
the
ability
of
the
liver
to
metabolize
cyanide
to
thiocyanate
by
first­
pass
metabolism.
By
contrast,

administration
of
the
same
dose
over
the
course
of
the
day
in
feed
or
water
allows
the
liver
to
fully
metabolize
the
cyanide.

In
light
of
the
uncertainty
regarding
the
actual
dose
administered
in
feed
by
Palmer
and
Olson
(
1979),
and
because
drinking
water
is
the
exposure
route
of
interest,
the
drinking
water
LOAEL
of
14
mg
CN/
kg­
day
(
Palmer
and
Olson,
1979)
is
used
as
the
basis
for
the
Ten­
day
Health
Advisory.
This
LOAEL
is
supported
by
the
BMDL
of
2.1
mg/
kg­
day
calculated
based
on
increased
plasma
TSH
in
protein­
and
iodine­
deficient
rats
(
a
sensitive
population)
in
the
Kreutler
et
al.
(
1978)
study.
Note
that
the
drinking
water
LOAEL
in
Palmer
and
Olson
(
1979)
is
approximately
a
factor
of
9
times
the
NOAEL
in
feed
in
this
study.
This
means
that
the
Health
Advisory
calculated
using
the
drinking
water
study
(
and
an
uncertainty
factor
of
10
for
extrapolating
from
a
LOAEL)
is
similar
to
the
value
that
would
be
calculated
using
the
NOAEL
in
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
12
Final
draft
feed,
but
the
latter
value
would
have
greater
uncertainty,
due
to
the
uncertainty
in
the
ingested
dose.
An
uncertainty
factor
of
10
is
used
for
animal­
to­
human
extrapolation,
an
uncertainty
factor
of
10
is
used
to
protect
sensitive
human
populations
(
i.
e.,
to
account
for
human
variability),
and
an
uncertainty
factor
of
10
is
used
to
account
for
extrapolation
from
a
LOAEL.

A
similar
Health
Advisory
could
be
calculated
from
the
BMDL
in
the
Kreutler
et
al.

(
1978)
study.
In
this
case,
no
factor
would
be
needed
for
extrapolation
from
a
LOAEL,
and
the
default
factor
of
10
would
be
used
to
extrapolate
from
animals
to
humans.
Since
an
effect
was
seen
only
in
the
rats
fed
a
protein­
deficient
diet,
the
study
was
done
in
a
sensitive
population,
and
no
extrapolation
would
be
needed
to
consider
toxicodynamic
variability
relating
to
a
sensitive
subpopulation.
Therefore
the
uncertainty
factor
for
human
variability
would
be
reduced
to
a
factor
of
3,
for
consideration
of
toxicokinetic
variability.
However,
in
light
of
the
large
degree
of
extrapolation
involved
and
the
uncertainty
related
to
the
BMDL
being
based
on
only
two
data
points,
confidence
in
the
BMDL
is
limited.
Therefore,
it
was
not
considered
appropriate
to
consider
the
Kreutler
et
al.
(
1978)
study
as
a
co­
critical
study.
As
shown
in
Table
VIII­
5:

(
14
mg/
kg­
day)
(
10
kg)
Ten­
day
HA
(
for
a
child)
=
=
0.14
mg/
L,
rounded
to
0.1
mg/
L
(
1000)
(
1
L/
day)
where:

14
mg/
kg­
day
=
LOAEL,
based
on
increased
liver
weight
in
rats
receiving
14
mg
CN/
kgday
for
21
days
in
drinking
water
(
Palmer
and
Olson,
1979)

10
kg
=
assumed
body
weight
of
a
child.

1000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
animals,
protection
of
sensitive
subpopulations,
and
extrapolation
from
a
LOAEL
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
13
Final
draft
1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

B.
1.3.
Longer­
term
Health
Advisory
for
Cyanide
The
most
appropriate
study
on
which
to
base
the
longer­
term
HA
is
the
subchronic
bioassay
by
NTP
(
1993),
which
evaluated
the
effects
of
several
doses
on
all
organ
systems
in
both
mice
and
rats.
The
study
identified
a
NOAEL
of
4.5
mg/
kg­
day
and
a
LOAEL
of
12.5
mg/
kg­
day
based
on
male
reproductive
effects
(
decreased
epididymal
weight,
testis
weight,
and
sperm
count)

in
rats.
Although
decreases
in
left
cauda­
epididymis
weight
and
sperm
motility
were
observed
at
doses
of
1.4
and
4.5
mg
CN/
kg­
day,
these
changes
were
insufficient
for
these
doses
to
be
considered
LOAELs.
NTP
(
1993)
did
not
consider
the
effects
on
sperm
motility
to
be
adverse
because
they
were
small
and
well
within
the
range
of
historical
controls.
Similarly,
the
decrease
in
left
cauda
epididymis
weight
was
small
at
1.4
and
4.5
mg/
kg­
day,
and
there
was
little
evidence
of
a
dose­
response
at
these
doses.
A
BMDL
of
0.79
mg/
kg­
day
was
calculated
for
this
study,
based
on
decreased
epididymal
weight,
and
supported
by
a
BMDL
of
1.3
mg/
kg­
day
for
both
decreased
spermatid
heads/
testis
and
decreased
spermatid
count.
No
studies
on
the
effects
of
cyanide
on
reproduction
function
in
males
have
been
located,
although
NTP
(
1993)
stated
that
comparison
with
other
studies
indicates
that
the
observed
decreases
would
not
result
in
functional
effects.

Two
other
subchronic
studies
identified
lower
NOAELs
and
LOAELs;
however,
these
studies
are
not
as
appropriate.
Jackson
(
1988)
identified
a
NOAEL
of
0.7
and
a
LOAEL
of
1.2
mg/
kg­
day
in
miniature
pigs
based
on
altered
thyroid
hormones
and
behavioral
changes.
This
NOAEL
is
not
considered
reliable,
however,
because
the
study
was
not
controlled
sufficiently
well
to
ensure
that
the
observed
differences
were
not
due
to
variability,
particularly
since
unequal
numbers
of
males
and
females
were
used
in
each
dose
group
and
the
observed
endpoints
tended
to
differ
between
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
14
Final
draft
the
two
sexes.
Kamalu
(
1993)
identified
a
LOAEL
of
1.04
mg/
kg­
day
in
dogs
based
on
male
reproductive
effects
and
histopathological
changes
in
the
kidney
and
adrenal
gland.
This
appears
to
be
a
well­
conducted
study,
but
dogs
are
not
a
suitable
model
for
the
toxicity
of
cyanide
in
humans,
because
their
levels
of
rhodanese
(
the
enzyme
which
detoxifies
cyanide)
are
much
lower
than
the
levels
in
humans
(
ATSDR,
1997).

The
database
for
cyanide
includes
a
subchronic
systemic
toxicity
study
in
rats
and
mice
(
NTP,
1993).
No
standard
multigeneration
or
developmental
toxicity
studies
of
cyanide
are
available.
The
single
available
developmental
toxicity
(
Tewe
and
Maner,
1981)
found
no
effects
on
litter
size,
birth
weight
of
pups,
or
pup
mortality
in
the
offspring
of
female
rats
receiving
cyanide
in
the
diet
throughout
mating,
gestation,
and
lactation,
although
decreased
pup
body
weight
gain
compared
to
the
low­
cyanide
control
was
seen
in
the
high­
cyanide
dose
(
weanling
dose)
of
34.3
mg/
kg­
day;
morphological
assessment
of
developmental
effects
was
not
conducted.

The
absence
of
a
study
evaluating
reproductive
function
is
of
particular
concern,
since
the
critical
effect
is
histopathological
changes
to
the
male
reproductive
organs
(
NTP,
1993).
In
addition,
the
absence
of
a
developmental
toxicity
study
that
evaluated
thyroid
hormone
or
neurodevelopmental
endpoints
is
of
concern,
since
prolonged
exposure
to
cyanide
is
known
to
affect
thyroid
hormone
levels
(
with
a
LOAEL
of
44
mg/
kg­
day
and
no
NOAEL
identified
in
Philbrick
et
al.,
1979),
and
neurodevelopmental
effects
are
known
to
result
from
decreased
T4
levels
(
Chan
and
Kilby,

2000).
As
discussed
in
Chapter
7,
there
is
not
a
direct
correlation
between
serum
levels
of
thyroid
hormones
and
neurodevelopmental
effects.
Therefore,
the
Philbrick
et
al.
(
1979)
data
are
inadequate
to
determine
what
doses
may
cause
developmental
neurotoxicity.
The
absence
of
data
on
doses
that
affect
male
reproductive
function,
and
particularly
the
absence
of
a
two­
generation
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
15
Final
draft
study,
warrants
at
least
a
factor
of
3
to
account
for
database
deficiencies.
The
absence
of
any
studies
on
neurobehavioral
or
neurodevelopmental
toxicity,
secondary
to
decreased
T4
levels
warrants
an
additional
factor
of
3
for
database
deficiencies.
Therefore,
a
total
database
uncertainty
factor
of
10
is
used.
No
quantitative
data
are
available
on
toxicokinetic
and
toxicodynamic
differences
between
animals
and
humans,
or
on
human
variability
in
these
areas.

Therefore,
the
default
uncertainty
factors
of
10
each
are
used
for
extrapolation
from
animals
to
humans,
and
for
protection
of
sensitive
individuals.
A
composite
uncertainty
factor
of
1000
results,
based
on
default
factors
of
10
each
for
interspecies
extrapolation
and
human
variability,

and
a
factor
of
10
for
database
uncertainties.
The
same
composite
uncertainty
factor
of
1000
was
used
for
the
derivation
of
a
Health
Advisory
based
on
the
BMDL.
No
additional
uncertainty
factor
is
needed
for
extrapolation
from
the
BMDL,
in
light
of
the
minimal
adversity
of
the
observed
effects,
and
since
the
BMDL
is
lower
than
the
corresponding
NOAEL.

Derivation
of
the
Longer­
term
HA
based
on
the
study
NOAEL
(
4.5
mg/
kg­
day)
(
10
kg)
Longer­
Term
HA
(
for
a
child)
=
=
0.045
mg/
L
(
rounded
to
0.05
mg/
L)
(
1000)
(
1
L/
day)

where:

4.5
mg/
kg­
day
=
NOAEL,
based
on
absence
of
effects
on
male
reproductive
organ
weight
and
spermatid
counts
in
rats
exposed
to
cyanide
in
drinking
water
for
13
weeks,
with
a
corresponding
LOAEL
of
12.5
mg/
kg­
day
(
NTP,
1993).

10
kg
=
assumed
body
weight
of
a
child.

1000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
NOAEL
in
animals,
protection
of
sensitive
individuals,
and
insufficiencies
in
the
database.
The
database
insufficiencies
include
the
absence
of
a
2­
generation
reproduction
study,
and
the
absence
of
adequate
developmental
toxicity
studies,
particularly
a
neurodevelopmental
toxicity
study,
in
light
of
the
potential
for
low­
level
cyanide
exposure
to
decrease
T4
levels.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
16
Final
draft
1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

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

(
4.5
mg/
kg­
day)
(
70
kg)
Longer­
Term
HA
(
for
adults)
=
=
0.158
mg/
L
(
rounded
to
0.2
mg/
L)
(
1000)(
2
L/
day)

where:

4.5
mg/
kg­
day
=
NOAEL,
based
on
absence
of
effects
on
male
reproductive
organ
weight
and
spermatid
counts
in
rats
exposed
to
cyanide
in
drinking
water
for
13
weeks,
with
a
corresponding
LOAEL
of
12.5
mg/
kg­
day
(
NTP,
1993).

70
kg
=
assumed
body
weight
of
an
adult.

1000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
NOAEL
in
animals,
inter­
individual
variability
in
humans,
and
insufficiencies
in
the
database.
The
database
insufficiencies
include
the
absence
of
a
2­
generation
reproduction
study,
and
the
absence
of
adequate
developmental
toxicity
studies,
particularly
a
neurodevelopmental
toxicity
study,
in
light
of
the
potential
for
low­
level
cyanide
exposure
to
decrease
T4
levels.

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

Derivation
of
the
Longer­
term
HA
based
on
the
study
BMDL
(
0.79
mg/
kg­
day)
(
10
kg)
Longer­
Term
HA
(
child)
=
=
0.0079
mg/
L
(
rounded
to
0.008
mg/
L)
(
1000)
(
1
L/
day)

where:

0.79
mg/
kg­
day
=
BMDL,
based
on
decreased
left
epididymis
weight
in
male
rats
exposed
to
cyanide
in
drinking
water
for
13
weeks
(
NTP,
1993).

10
kg
=
assumed
body
weight
of
a
child.

1000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
BMDL
in
animals,
protection
of
sensitive
individuals,
and
insufficiencies
in
the
database.
The
database
insufficiencies
include
the
absence
of
a
2­
generation
reproduction
study,
and
the
absence
of
adequate
developmental
toxicity
studies,
particularly
a
neurodevelopmental
toxicity
study,
in
light
of
the
potential
for
low­
level
cyanide
exposure
to
decrease
T4
levels.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
1An
uncertainty
factor
of
10
each
was
used
for
extrapolation
from
animals
to
humans,
and
for
protection
of
sensitive
human
populations.
A
modifying
factor
of
5
was
used
to
account
for
the
apparent
tolerance
to
cyanide
when
it
is
ingested
with
food
rather
than
when
it
is
administered
by
gavage
or
by
EPA/
OW/
OST/
HECD
VIII­
17
Final
draft
1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

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

(
0.79
mg/
kg­
day)
(
70
kg)
Longer­
Term
HA
(
for
adults)
=
=
0.028
mg/
L
(
rounded
to
0.03
mg/
L)
(
1000)(
2
L/
day)

where:

0.79
mg/
kg­
day
=
BMDL,
based
on
decreased
left
epididymis
weight
in
male
rats
exposed
to
cyanide
in
drinking
water
for
13
weeks
(
NTP,
1993).

70
kg
=
assumed
body
weight
of
an
adult.

1000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
NOAEL
in
animals,
inter­
individual
variability
in
humans,
and
insufficiencies
in
the
database.
The
database
insufficiencies
include
the
absence
of
a
2­
generation
reproduction
study,
and
the
absence
of
adequate
developmental
toxicity
studies,
particularly
a
neurodevelopmental
toxicity
study,
in
light
of
the
potential
for
low­
level
cyanide
exposure
to
decrease
T4
levels.

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

B.
1.4
Reference
Dose,
Drinking
Water
Equivalent
Level,
and
Lifetime
Health
Advisory
for
Cyanide
The
only
chronic
oral
study
of
cyanide
(
Howard
and
Hanzal,
1955)
identified
a
freestanding
NOAEL
of
10.8
mg/
kg­
day.
This
study,
together
with
the
dietary
study
of
Philbrick
et
al.
(
1979),
which
identified
a
LOAEL
of
44
mg/
kg­
day
based
on
myelin
degeneration
and
increased
thyroid
weight,
were
used
by
EPA
in
1985
to
develop
an
RfD
for
free
cyanide
of
0.02
mg/
kg­
day,
and
an
RfD
for
cyanogen
chloride
of
0.05
mg/
kg­
day
(
as
reported
in
U.
S.
EPA,

2001a).
1
However,
the
more
recent
subchronic
bioassay
by
NTP
(
1993)
is
a
better­
quality
study
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
drinking
water.
These
RfDs
were
developed
prior
to
EPA's
adoption
of
the
database
uncertainty
factor.
In
addition,
a
default
food
factor
of
0.05
was
used
to
calculate
the
dose
in
the
Philbrick
et
al.
(
1979)
study,
leading
to
a
LOAEL
of
30
mg/
kg­
day.

EPA/
OW/
OST/
HECD
VIII­
18
Final
draft
with
less
uncertainty,
because
it
identifies
both
a
NOAEL
and
a
LOAEL,
and,
therefore,
is
more
appropriate
as
the
basis
of
the
RfD.
ATSDR
(
1997)
used
the
NTP
(
1993)
study
as
the
basis
of
an
intermediate
oral
minimal­
risk
level
(
MRL)
for
cyanide.
ATSDR
determined
that
the
high
dose
in
male
rats,
12.5
mg
CN/
kg­
day,
was
the
LOAEL
based
on
the
spectrum
of
reproductive
effects
observed
at
this
dose.
The
next­
lowest
dose,
4.5
mg
CN/
kg­
day,
was
designated
as
the
NOAEL.

Although
decreases
in
left
cauda­
epididymis
weight
and
sperm
motility
were
observed
at
doses
of
1.4
and
4.5
mg
CN/
kg­
day,
NTP
(
1993)
(
and
ATSDR)
determined
that
these
effects
by
themselves
were
not
adverse.
A
total
uncertainty
factor
of
100
was
applied,
a
factor
of
10
to
extrapolate
from
animals
to
humans,
and
another
factor
of
10
to
account
for
human
variability.

The
resulting
value
for
the
intermediate
MRL
is
0.05
mg/
kg­
day.

As
discussed
in
the
context
of
the
Longer­
term
Health
Advisory,
NTP
(
1993)
identified
a
NOAEL
of
4.5
mg/
kg­
day
and
a
LOAEL
of
12.5
mg/
kg­
day
based
on
male
reproductive
effects
(
decreased
epididymal
weight,
testis
weight,
and
sperm
count).
The
effects
on
sperm
motility
at
lower
doses
were
not
considered
adverse
because
the
changes
were
small
and
well
within
the
range
of
historical
controls.
Similarly,
the
decrease
in
left
cauda
epididymis
weight
was
small
at
1.4
and
4.5
mg/
kg­
day,
and
there
was
little
evidence
of
a
dose­
response
at
these
doses.
A
BMDL
of
0.79
mg/
kg­
day
was
calculated
for
this
study,
based
on
decreased
left
epididymis
weight,
and
supported
by
a
BMDL
of
1.3
mg/
kg­
day
for
both
decreased
spermatid
heads/
testis
and
decreased
spermatid
count.
No
studies
on
the
effects
of
cyanide
on
reproductive
function
in
males
have
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
19
Final
draft
been
located,
although
NTP
(
1993)
stated
that
comparison
with
other
studies
indicates
that
the
observed
decreases
would
not
result
in
functional
effects.
For
development
of
the
Lifetime
Health
Advisory,
full
uncertainty
factors
of
10
were
used
to
account
for
extrapolation
from
animals
and
to
account
for
human
variability.
The
database
for
cyanide
includes
systemic
toxicity
studies
in
rats
and
mice,
as
well
as
lower­
quality
studies
in
pigs
and
dogs.
However,
no
standard
multigeneration
or
developmental
toxicity
studies
of
cyanide
are
available.
In
the
only
available
study
on
developmental
effects
of
cyanide
(
Tewe
and
Maner,
1981)
the
only
observed
effect
was
that
pup
body
weight
gain
was
decreased
at
the
high
dose
of
34.3
mg/
kg­
day
(
weanling
dose)

compared
to
the
low
"
control"
cyanide
dose;
morphological
assessment
of
developmental
effects
was
not
conducted.
As
noted
above,
the
absence
of
a
study
evaluating
reproductive
function
and
of
a
developmental
toxicity
study
(
particularly
the
absence
of
a
neurodevelopmental
study)
are
of
particular
concern,
since
the
current
critical
effect
is
histopathological
changes
to
the
male
reproductive
organs
(
NTP,
1993).
In
addition,
cyanide
is
known
to
affect
thyroid
hormone
levels,

and
agents
that
affect
thyroid
hormone
levels
may
also
affect
neurological
development.
Based
on
the
lack
of
a
multigeneration
reproductive
toxicity
study
and
concerns
about
potential
neurodevelopmental
toxicity,
a
factor
of
10
was
used
for
database
uncertainties
for
the
Longerterm
cyanide
Health
Advisory.

The
data
indicate
that
a
reduced
uncertainty
factor
for
extrapolation
from
subchronic
to
chronic
duration
is
appropriate.
There
was
no
progression
in
liver
effects
(
actually,
a
regression)

with
increased
exposure.
Palmer
and
Olson
(
1979)
found
a
LOAEL
of
12
mg
CN/
kg­
day
for
increased
liver
weight
in
male
rats
exposed
via
a
drinking­
water
study
for
21
days,
but
there
was
no
effect
on
liver
weight
in
male
rats
exposed
via
drinking
water
for
13
weeks
(
NTP,
1993).
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
20
Final
draft
There
was
an
increase
in
liver
weight
in
female
rats
in
the
NTP
(
1993)
study,
but
this
increase
was
comparable
to
the
increase
seen
in
males
in
the
short­
term
study;
Palmer
and
Olson
(
1979)
did
not
evaluate
female
rats.
As
discussed
in
the
context
of
the
Longer­
term
Health
Advisory
for
thiocyanate,
progression
of
effect
from
subchronic
to
chronic
exposure
would
not
be
expected
for
effects
of
cyanide
that
are
secondary
to
thyroid
hormone
changes,
since
the
thyroid
responds
rapidly
to
iodine­
like
ions,
and
thiocyanate
does
not
accumulate
in
the
body.
However,
the
mode
of
action
of
the
critical
effect
on
the
male
reproductive
tract
has
not
been
identified
(
e.
g.,
whether
it
is
due
to
cyanide
or
thiocyanate,
and
whether
it
is
secondary
to
effects
on
thyroid
hormones).
It
is
also
not
known
whether
the
observed
effects
would
progress
with
continued
exposure.

Therefore,
the
default
factor
of
10
is
used
for
subchronic­
to­
chronic
extrapolation
for
the
male
reproductive
effects.

An
endpoint­
by­
endpoint
analysis
is
needed,
however,
for
determination
of
the
appropriate
composite
uncertainty
factor
for
subchronic
to
chronic
extrapolation
and
for
database
uncertainties.
If
the
"
true"
critical
effect
is
decreased
weight
of
male
reproductive
organs
(
as
in
the
NTP,
1993
study),
a
factor
of
10
would
be
needed
for
subchronic
to
chronic
extrapolation,

but
no
database
uncertainty
factor
would
be
needed
(
since
the
critical
effect
has
been
identified).

If
neurodevelopmental
effects
due
to
alterations
in
thyroid
hormones
are
the
"
true"
critical
effect,

a
full
database
uncertainty
factor
of
10
would
be
needed,
but
a
factor
of
1
would
be
adequate
for
subchronic
to
chronic
extrapolation,
as
described
in
the
context
of
the
Longer­
term
health
advisory
for
thiocyanate.
Similarly,
if
the
"
true"
critical
effect
were
a
functional
effect
identified
in
a
2­
generation
reproduction
study,
a
database
factor
of
3
would
be
adequate,
based
on
the
usual
approach
in
the
absence
of
a
multigeneration
study.
No
subchronic
to
chronic
factor
would
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
21
Final
draft
be
needed,
because
a
subchronic
to
chronic
uncertainty
factor
is
not
usually
used
with
a
multigeneration
study.
Therefore,
regardless
of
the
"
true"
critical
effect,
a
composite
uncertainty
factor
of
10
is
adequate
for
database
insufficiencies
and
for
subchronic
to
chronic
extrapolation.

Thus,
full
factors
of
10
were
used
for
extrapolation
from
animals
to
humans
and
for
human
variability,
and
composite
factor
of
10
was
used
for
database
uncertainties
and
for
subchronic
to
chronic
extrapolation.
A
composite
uncertainty
factor
of
1000
results.
Because
the
principal
study
(
NTP,
1993)
used
drinking­
water
administration,
no
additional
modifying
factor
is
needed.
The
same
composite
uncertainty
factor
of
1000
would
be
used
with
the
BMDL
of
0.79
mg/
kg­
day,
in
light
of
the
sensitivity
of
the
critical
effect
(
decreased
left
epididymis
weight),
and
because
the
BMDL
is
markedly
lower
than
the
NOAEL.

Derivation
of
the
Lifetime
HA
based
on
the
study
NOAEL
Step
1:
Determination
of
a
RfD
for
Cyanide
RfD
=
4.5
mg/
kg­
day
=
0.0045
mg/
kg­
day
(
rounded
to
0.005
mg/
kg­
day)
1000
where:

4.5
mg/
kg­
day
=
NOAEL,
based
on
absence
of
effects
on
male
reproductive
organ
weight
(
epididymis,
cauda
epididymis,
and
testes
weight)
and
spermatid
counts
in
rats
exposed
to
cyanide
in
drinking
water
for
13
weeks,
with
a
corresponding
LOAEL
of
12.5
mg/
kg­
day
(
NTP,
1993).

1000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
NOAEL
in
animals,
protection
of
sensitive
subpopulations,
and
insufficiencies
in
the
database.
The
database
insufficiencies
include
the
absence
of
a
2­
generation
reproduction
study,
and
the
absence
of
adequate
developmental
toxicity
studies,
particularly
a
neurodevelopmental
toxicity
study,
in
light
of
the
potential
for
low­
level
cyanide
exposure
to
decrease
T4
levels.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
22
Final
draft
Step
2:
Determination
of
a
Drinking
Water
Equivalent
Level
(
DWEL)
for
Cyanide
DWEL
=
(
0.0045
mg/
kg­
day)
(
70kg)
=
0.16
mg/
L
(
rounded
to
0.2
mg/
L)
(
2
L/
day)

where:

0.0045
mg/
kg­
day
=
RfD
(
before
rounding)
70
kg
=
assumed
body
weight
of
an
adult
2
L/
day
=
assumed
drinking­
water
consumption
of
a
70­
kg
adult
Step
3:
Determination
of
Lifetime
HA
for
Cyanide
Lifetime
HA
=
(
0.16
mg/
L)
(
20%)
=
0.032
mg/
L
(
rounded
to
30

g/
L)

where:
0.16
mg/
L
=
DWEL
20%
=
assumed
relative
source
contribution
from
water
Derivation
of
the
Lifetime
HA
based
on
the
study
BMDL
Step
1:
Determination
of
a
RfD
for
Cyanide
RfD
=
0.79
mg/
kg­
day
=
0.00079
mg/
kg­
day
(
rounded
to
0.0008
mg/
kg­
day)
1000
where:

0.79
mg/
kg­
day
=
BMDL,
based
on
decreased
left
epididymis
weight
in
male
rats
exposed
to
cyanide
in
drinking
water
for
13
weeks
(
NTP,
1993).

1000
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
BMDL
in
animals,
protection
of
sensitive
subpopulations,
and
insufficiencies
in
the
database.
The
database
insufficiencies
include
the
absence
of
a
2­
generation
reproduction
study,
and
the
absence
of
adequate
developmental
toxicity
studies,
particularly
a
neurodevelopmental
toxicity
study,
in
light
of
the
potential
for
low­
level
cyanide
exposure
to
decrease
T4
levels.

Step
2:
Determination
of
a
Drinking
Water
Equivalent
Level
(
DWEL)
for
Cyanide
DWEL
=
(
0.00079
mg/
kg­
day)
(
70kg)
=
0.028
mg/
L
(
rounded
to
0.03
mg/
L)
(
2
L/
day)
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
23
Final
draft
where:

0.00079
mg/
kg­
day
=
RfD
(
before
rounding)
70
kg
=
assumed
body
weight
of
an
adult
2
L/
day
=
assumed
drinking­
water
consumption
of
a
70­
kg
adult
Step
3:
Determination
of
Lifetime
HA
for
Cyanide
Lifetime
HA
=
(
0.028
mg/
L)
(
20%)
=
0.0056
mg/
L
(
rounded
to
6

g/
L)

where:
0.028
mg/
L
=
DWEL
20%
=
assumed
relative
source
contribution
from
water
B.
2
Thiocyanate
B.
2.1
One­
day
Health
Advisory
for
Thiocyanate
No
suitable
studies
were
located.
Decreased
hemoglobin
was
observed
in
guinea
pigs
following
2
doses
of
120
mg
SCN/
kg/
dose,
the
only
dose
tested
(
Taubman
and
Heilborn,
1930),

but
this
study
is
not
of
adequate
quality
to
use
as
the
basis
for
the
development
of
a
Health
Advisory.
In
the
absence
of
adequate
data,
the
ten­
day
HA
value
is
recommended
as
a
conservative
estimate
of
an
appropriate
one­
day
HA
value.

B.
2.2
Ten­
day
Health
Advisory
for
Thiocyanate
No
NOAEL
was
identified
in
the
available
short­
term
animal
toxicity
studies
(
Wolff
et
al.,

1946;
Rawson
et
al.,
1944;
De
Groot
et
al.,
1991).
The
lowest
reliable
LOAEL
was
123
mg
SCN/
kg­
day,
based
on
decreased
T4,
increased
thyroid
weight,
and
activation
of
thyroid
follicles
(
De
Groot
et
al.,
1991);
no
NOAEL
was
identified
in
this
study.
Smith
and
Rudolf
(
1928)

demonstrated
a
15­
30
mm
Hg
drop
in
systolic
blood
pressure
in
normotensive
human
subjects
treated
with
three
daily
doses
of
3.3
mg
SCN/
kg­
day
(
for
a
cumulative
dose
of
10
mg/
kg­
day)
for
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
24
Final
draft
1
week,
with
no
other
effects
reported.
However,
the
likely
target
organ
(
the
thyroid)
was
not
evaluated
in
this
study.
A
human
LOAEL
of
1.5­
1.7
mg
SCN/
kg­
day
was
identified
by
Palmer
et
al.
(
1929)
and
Palmer
and
Sprague
(
1929),
based
on
weakness,
angina,
and
precordial
distress
in
hypertensive
subjects
administered
the
TWA
dose
over
3­
4
weeks
(
decreasing
daily
doses).
It
was
unclear,
however,
whether
the
observed
effects
were
secondary
to
a
rapid
decrease
in
blood
pressure.
Thyroid
effects
were
not
evaluated
by
Palmer
and
colleagues.
However,
the
LOAEL
in
the
hypertensive
subjects
is
consistent
with
the
NOAEL
of
0.11
mg
SCN/
kg­
day
identified
by
Dahlberg
et
al.
(
1984),
based
on
the
absence
of
changes
in
T3,
T4,
or
thyrotropic
hormone
in
normotensive
subjects
treated
for
12
weeks.
Higher
doses
were
not
tested
by
Dahlberg
et
al.

(
1984).
The
results
of
Dahlberg
et
al.
(
1984)
indicate
that
the
NOAEL
for
a
ten­
day
exposure
would
be
0.11
mg
SCN/
kg­
day
or
higher,
and
that
the
thyroid
is
adequately
protected
by
deriving
the
Ten­
day
Health
advisory
from
a
LOAEL
of
1.5
mg
SCN/
kg­
day,
and
using
an
uncertainty
factor
of
10
to
extrapolate
to
a
NOAEL.
Thus,
the
Ten­
day
Health
Advisory
is
based
on
the
LOAEL
of
1.5
mg
SCN/
kg­
day
identified
by
Palmer
et
al.
(
1929)
and
Palmer
and
Sprague
(
1929)

in
hypertensive
subjects.
An
uncertainty
factor
of
10
is
used
to
protect
sensitive
subpopulations,

and
an
uncertainty
factor
of
10
is
used
to
account
for
the
use
of
a
LOAEL,
resulting
in
a
composite
uncertainty
factor
of
100.
(
See
Table
VIII­
6.)

(
1.5
mg/
kg­
day)
(
10
kg)
Ten­
day
HA
(
for
a
child)
=
=
0.15
mg/
L,
rounded
to
0.2
mg/
L
(
100)(
1
L/
day)

where:

1.5
mg/
kg­
day
=
LOAEL,
at
which
weakness,
angina,
and
precordial
distress
were
seen
in
hypertensive
subjects
administered
the
TWA
dose
for
3­
4
weeks
(
Palmer
et
al.,
1929;
Palmer
and
Sprague,
1929)
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
25
Final
draft
10
kg
=
assumed
body
weight
of
a
child.

100
=
composite
uncertainty
factor,
chosen
to
account
for
extrapolation
from
a
LOAEL
in
humans
and
protection
of
sensitive
subpopulations
1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

B.
2.3
Longer­
term
Health
Advisory
for
Thiocyanate
Long­
term
toxicity
data
are
available
from
mice
(
Nagasawa
et
al.,
1980)
and
rats
(
Pyska,

1977;
Philbrick
et
al.,
1979;
Kanno
et
al.,
1990)
exposed
to
thiocyanate
via
the
diet
and
drinking
water.
These
data
support
the
identification
of
thyroid
as
the
target
organ.
None
of
the
studies
that
evaluated
thyroid
endpoints
identified
a
NOAEL.
The
lowest
LOAEL
was
identified
by
Philbrick
et
al.
(
1979),
who
reported
a
statistically
significant
decrease
in
plasma
T4
and
increased
thyroid
weight
(
but
no
thyroid
histopathology)
in
rats
administered
67
mg
SCN/
kg­
day
in
diet
for
11.5
months;
plasma
T4
was
increased
at
this
dose
after
4
months
of
exposure.
Pyska
et
al.

(
1977)
also
reported
decreased
plasma
protein­
bound
iodine
(
a
measure
of
plasma
T3
and
T4)
in
rats
administered
90
mg
SCN/
kg­
day
for
approximately
10
weeks
in
drinking
water.
No
standard
multigeneration
reproduction
or
developmental
toxicity
study
of
thiocyanate
by
any
route
was
located.
However,
a
multigeneration
study
that
evaluated
thyroid
hormone
levels
in
rat
pups
is
available
(
Raghunath
and
Bala,
1998),
as
well
as
two
developmental
toxicity
studies
that
evaluated
thyroid
effects
in
rat
pups
(
Bala
et
al.,
1996;
Kreutler
et
al.,
1978).
These
studies
have
found
effects
on
maternal
thyroid
hormones
at
the
lowest
dose
evaluated
in
dams
(
52
mg
SCN/
kg­
day;
Pyska,
1977)
and
in
pups
(
108
mg
SCN/
kg­
day,
Raghunath
and
Bala,
1998).
A
mouse
reproductive
toxicity
study
found
no
effect
on
pup
weight
up
to
341
mg
SCN/
kg­
day,
but
did
not
evaluate
pup
morphology
or
thyroid
endpoints
.
Kreutler
et
al.
(
1978)
found
increased
relative
thyroid
weight
in
pups
born
to
the
lowest
dose
group,
5.6
mg
SCN/
kg­
day
in
drinking
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
26
Final
draft
water.
Unfortunately,
thyroid
hormone
levels
and
histopathology
were
not
evaluated,
and
increases
in
thyroid
weight
may
be
adaptive
or
adverse,
depending
on
the
alterations
in
thyroid
hormones
and
histopathology.

Despite
the
availability
of
LOAELs
(
without
corresponding
NOAELs)
for
thyroid
hormone
effects,
one
can
not
confidently
conclude
that
the
NOAEL
for
neurodevelopmental
effects
is
known
within
a
factor
of
ten
(
the
uncertainty
factor
for
extrapolating
from
a
LOAEL
to
a
NOAEL)
of
the
LOAEL
for
thyroid
hormone
changes
in
pups.
There
are
several
reasons
for
this
uncertainty.
First,
the
degree
of
decrease
in
serum
T4
that
results
in
neurodevelopmental
effects
has
not
been
determined.
In
addition,
thyroid
hormones
are
controlled
by
a
negative
feedback
loop,
and
rats
rapidly
up­
regulate
serum
thyroid
hormone
levels.
Because
of
this
rapid
feedback,
the
peak
degree
of
change
in
hormone
levels
in
the
developmental
toxicity
studies
could
have
been
much
larger
at
earlier
time
points.
Since
developmental
toxicity
depends
on
exposure
during
a
critical
developmental
window,
a
large,
but
brief,
change
in
thyroid
hormone
levels
might
result
in
marked
neurodevelopmental
effects.
This
means
that
the
ratio
between
the
free­
standing
LOAELs
identified
for
changes
in
thyroid
hormone
levels
in
rat
pups
and
the
short­
term
NOAEL
(
in
the
same
species)
for
this
effect
may
be
more
than
an
a
factor
of
ten.
Because
a
short­
term
decrease
in
T4
levels
(
of
sufficient
magnitude)
can
affect
neurodevelopment
if
the
change
occurs
during
the
window
of
sensitivity
at
the
appropriate
stage
in
development,
even
short­
term
deviations
in
T4
are
of
concern.
Although
the
increased
relative
thyroid
weight
observed
by
Kreutler
et
al.
(
1978)
at
5.6
mg/
kg­
day
may
have
been
adaptive,
that
observation
(
at
a
dose
20­

fold
below
the
LOAEL
for
thyroid
hormone
changes
observed
by
Bala
et
al.
[
1996])
suggests
that
there
may
have
been
a
short­
term
change
in
thyroid
hormone
levels
at
lower
doses,
to
which
the
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
27
Final
draft
rats
later
adapted.
Based
on
these
considerations,
a
neurodevelopmental
toxicity
study
would
be
important
in
determining
whether
the
brain
is
affected
at
doses
below
those
observed
to
affect
serum
hormone
levels
in
rat
pups.

Human
studies
provide
a
reliable
apparent
NOAEL/
LOAEL
pair,
based
on
thyroid
effects
in
normotensive
populations.
Dahlberg
et
al.
(
1984)
found
no
effect
on
serum
T3,
T4,
or
thyrotropic
hormone,
or
the
T3:
T4
ratio
in
37
volunteers
administered
8
mg/
day
thiocyanate
in
milk
(
0.11
mg
SCN/
kg­
day)
for
12
weeks.
Banerjee
et
al.
(
1997)
found
decreased
serum
T4
and
increased
TSH
compared
to
matched
controls
in
35
women
in
India
who
ingested
thiocyanate
and
hydrogen
peroxide
as
a
bacteriocide
in
milk
for
at
least
5
years.
The
exposed
women
ingested
approximately
0.19
mg
SCN/
kg­
day.
This
LOAEL
is
approximately
a
factor
of
2
higher
than
the
NOAEL,
suggesting
that
the
NOAEL/
LOAEL
boundary
is
well­
defined.
The
LOAEL
identified
by
Banerjee
et
al.
(
1997)
is
also
supported
by
the
results
of
Beamish
et
al.
(
1954),
who
reported
decreased
protein­
bound
plasma
iodine
and
decreased
thyroid
uptake
of
iodine
in
a
clinical
study
of
5
hypertensive
adults
in
Canada.
Although
doses
were
not
reported
in
the
Beamish
et
al.

(
1954)
study,
the
study
subjects
had
plasma
thiocyanate
levels
of
1.3­
5
mg/
100
mL,

concentrations
very
similar
to
the
average
blood
thiocyanate
level
of
1.3
mg/
100
mL
reported
by
Banerjee
et
al.
(
1997).
This
study
was
not
considered
co­
critical,
because
actual
intake
was
not
reported.
Note
that
the
Indian
population
might
constitute
a
sensitive
population
if
the
women
had
borderline
protein
intake.
There
is,
however,
no
support
for
this
hypothesis,
other
than
the
high
percentage
of
the
Indian
population
that
is
malnourished.
Since
the
study
subjects
were
from
Calcutta,
a
coastal
city,
it
appears
that
an
iodine
deficiency
was
unlikely.
In
the
absence
of
definitive
information
on
a
protein
or
iodine
deficiency,
the
study
population
is
conservatively
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
28
Final
draft
assumed
not
to
constitute
a
sensitive
population.
Although
the
NOAEL
is
from
a
study
of
shorter
duration,
the
rapid
nature
of
the
thyroid
response
to
iodine­
like
ions,
and
the
lack
of
accumulation
of
thiocyanate
in
the
body,
indicate
that
the
12­
week
NOAEL
is
adequately
protective
for
longerterm
exposure.
Therefore,
progression
of
the
effect
with
increased
exposure
duration
from
12
weeks
to
5
years
would
not
be
expected.

Using
the
NOAEL
of
0.11
mg
SCN/
kg­
day
from
Dahlberg
et
al.
(
1984),
an
uncertainty
factor
of
10
is
applied
to
protect
sensitive
subpopulations.
A
factor
of
10
is
applied
for
database
deficiencies.
A
factor
of
3
would
normally
be
applied
to
account
for
the
absence
of
a
standard
multi­
generation
study.
This
default
was
increased
to
10
because
a
NOAEL
for
effects
on
the
thyroid
in
developing
fetuses
has
not
been
identified;
effects
on
the
fetal
thyroid
are
of
particular
concern,
because
neurodevelopmental
effects
are
known
to
result
from
decreased
T4
levels
(
Chan
and
Kilby,
2000),
and
no
neurological
effects
were
evaluated
in
pups
exposed
to
thiocyanate
during
gestation.
The
composite
uncertainty
factor
is
100.

(
0.11
mg/
kg­
day)
(
10
kg)
Longer­
Term
HA
(
for
a
child)
=
=
0.011
mg/
L
(
rounded
to
0.01
mg/
L)
(
100)(
1
L/
day)
where:

0.11
mg/
kg­
day
=
NOAEL,
based
on
absence
of
effects
on
hormone
levels
in
human
subjects
exposed
to
thiocyanate
for
12
weeks
(
Dahlberg
et
al.,
1984)

10
kg
=
assumed
body
weight
of
a
child.

100
=
composite
uncertainty
factor,
chosen
to
account
for
inter­
individual
variability
in
humans,
and
insufficiencies
in
the
database,
including
the
absence
of
a
2­
generation
reproductive
study,
and
the
lack
of
adequate
information
on
effects
on
the
fetal
thyroid
effects
and
associated
neurodevelopmental
effects
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
29
Final
draft
1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.

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

(
0.11
mg/
kg­
day)
(
70
kg)
Longer­
Term
HA
(
for
adults)
=
=
0.039
mg/
L
(
rounded
to
0.04
mg/
L)
(
100)(
2
L/
day)
where:

0.11
mg/
kg­
day
=
NOAEL,
based
on
absence
of
effects
on
thyroid
hormone
levels
in
human
subjects
exposed
to
thiocyanate
for
12
weeks
(
Dahlberg
et
al.,
1984).

70
kg
=
assumed
body
weight
of
an
adult.

100
=
composite
uncertainty
factor,
chosen
to
account
for
inter­
individual
variability
in
humans
and
insufficiencies
in
the
database,
including
the
absence
of
a
2­
generation
reproductive
study,
and
the
lack
of
adequate
information
on
effects
on
the
fetal
thyroid
effects
and
associated
neurodevelopmental
effects.

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

B.
2.4
Reference
Dose,
Drinking
Water
Equivalent
Level,
and
Lifetime
Health
Advisory
for
Thiocyanate
The
only
chronic
animal
studies
of
thiocyanate
did
not
evaluate
sensitive
endpoints,
such
as
effects
on
thyroid
hormones.
As
discussed
in
the
context
of
the
Longer­
term
Health
Advisory,

the
thyroid
responds
rapidly
to
effects
of
ions
such
as
thiocyanate.
Therefore,
the
human
NOAEL
of
0.11
mg
SCN/
kg­
day,
based
on
a
12­
week
study
(
Dahlberg
et
al.,
1984),
coupled
with
a
human
LOAEL
of
0.19
mg
SCN/
kg­
day,
based
on
a
5­
year
study
(
Banerjee
et
al.,
1997),
is
appropriate
as
the
basis
for
the
Lifetime
health
advisory,
with
no
further
adjustment.
The
LOAEL
identified
by
Beamish
et
al.
(
1954)
also
supports
the
identified
critical
effect
level.
However,
Beamish
et
al.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
30
Final
draft
(
1954)
is
not
appropriate
as
a
co­
critical
study,
because
only
serum
thiocyanate
levels
were
reported,
without
any
information
on
intake.

As
noted
for
the
Longer­
term
Health
Advisory,
two
uncertainty
factors
are
used
with
this
NOAEL.
A
factor
of
10
is
applied
to
protect
sensitive
subpopulations,
and
a
factor
of
10
is
applied
for
database
deficiencies,
including
the
absence
of
a
2­
generation
reproductive
study
and
the
absence
of
a
neurodevelopmental
toxicity
study,
in
light
of
the
neurological
effects
of
decreased
T4
levels
in
the
developing
animal.
No
uncertainty
factor
is
necessary
for
subchronicto
chronic
extrapolation,
in
light
of
the
rapid
response
of
the
thyroid.

Step
1:
Determination
of
RfD
for
Thiocyanate
(
0.11
mg/
kg­
day)
RfD
=
=
0.0011
mg/
kg­
day,
rounded
to
0.001
mg/
kg­
day
(
100)

where:

0.11
mg/
kg­
day
=
NOAEL,
based
on
absence
of
effects
on
thyroid
in
human
subjects
exposed
to
thiocyanate
for
12
weeks
(
Dahlberg
et
al.,
1984).

100
=
composite
uncertainty
factor
chosen
to
protect
sensitive
subpopulations
and
to
account
for
insufficiencies
in
the
database,
including
the
absence
of
a
2­
generation
reproductive
study,
and
the
lack
of
adequate
information
on
effects
on
the
fetal
thyroid
effects
and
associated
neurodevelopmental
effects.

Step
2:
Determination
of
the
Drinking
Water
Equivalent
Level
(
DWEL)
for
Thiocyanate
(
0.0011
mg/
kg­
day)
(
70
kg)
DWEL
=
=
0.039
mg/
L
(
rounded
to
0.04
mg/
L)
(
2
L/
day)

where:

0.0011
mg/
kg­
day
=
RfD
(
before
rounding)

70
kg
=
assumed
body
weight
of
an
adult.
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
31
Final
draft
2
L/
day
=
assumed
water
consumption
of
a
70­
kg
adult.

Step
3:
Determination
of
Lifetime
HA
for
Thiocyanate
Lifetime
HA
=
(
0.039
mg/
L)
(
20%)
=
0.0078
mg/
L
(
rounded
to
8
µ
g/
L)

where:

0.039
mg/
L
=
DWEL
(
before
rounding)

20%
=
assumed
relative
source
contribution
from
water
B.
3
Cyanogen
chloride.

No
suitable
studies
are
available
for
developing
quantitative
assessments
of
cyanogen
chloride.
No
oral
studies
of
cyanogen
chloride
toxicity
are
available.
The
available
studies
(
Reed,

1920;
Haymaker,
1952;
Aldridge
and
Evans,
1946;
Flury
and
Zernik,
1931)
were
done
prior
to
the
development
of
modern
toxicology
methods,
used
inadequate
numbers
of
animals,
did
not
evaluate
standard
endpoints,
and
used
either
inhalation
or
injection
exposure.
In
the
absence
of
adequate
toxicity
data
on
cyanogen
chloride
itself,
an
alternative
approach
to
dose­
response
assessment
for
cyanogen
chloride
is
to
use
the
quantitation
for
cyanogen
chloride
metabolites
as
a
surrogate
for
cyanogen
chloride.
Early
studies
demonstrate
that
cyanide
and
thiocyanate
are
metabolites
of
cyanogen
chloride.
Cyanide
accounts
for
approximately
30­
40%
of
the
cyanogen
chloride
dose
at
high
doses,
and
60­
80%
of
the
cyanogen
chloride
dose
at
lower
doses
(
Aldridge
and
Evans,
1946;
Midwest
Research
Institute,
1997).
Because
thiocyanate
was
not
measured
and
a
mass
balance
was
not
calculated,
actual
conversion
to
cyanide
may
have
been
higher,
since
some
of
the
cyanide
would
have
been
converted
to
thiocyanate.
Conversion
to
cyanide
would
be
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
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Potential
Metabolites
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OST/
HECD
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32
Final
draft
expected
to
be
on
the
higher
end
(
i.
e.,
closer
to
the
80%
estimate)
at
environmentally­
relevant
doses.
As
discussed
in
Chapter
3,
there
are
uncertainties
in
this
conclusion,
in
light
of
the
absence
of
quantitative
mass
balance
data
on
cyanogen
chloride
metabolism.
For
example,
it
is
not
known
whether
cyanide
and
thiocyanate
together
account
for
100%
of
the
cyanogen
chloride
dose,
or
if
there
are
additional
metabolites
of
cyanogen
chloride
that
have
yet
to
be
identified.
As
discussed
in
Chapter
3,
the
aqueous
chemistry
of
cyanogen
chloride
indicates
that
other
potential
metabolites
are
cyanate,
cyanamide,
and
HCl
(
Price
et
al.,
1947;
Migridichian,
1946);
these
chemicals
are
discussed
in
Appendices
C,
D,
and
E,
respectively.
HCl
production
has
not
been
quantified
in
vivo
or
in
vitro,
but
mass
balance
considerations
mean
that
metabolism
of
cyanogen
chloride
would
produce
a
mole
of
HCl
for
every
mole
of
cyanogen
chloride.

As
discussed
in
Chapters
3
and
7,
ingested
cyanogen
chloride
may
react
with
nucleophiles
such
as
proteins
in
the
gastrointestinal
tract
contents;
cyanate
and
cyanamide
might
also
be
formed
slowly
in
the
gastrointestinal
tract,
but
hydrolysis
would
be
expected
to
be
very
slow,
due
to
the
low
pH
in
the
stomach.
Therefore
it
is
plausible
that
absorption
of
cyanogen
chloride
is
primarily
as
the
parent
form.
Absorbed
cyanogen
chloride
(
or
its
metabolites)
would
enter
the
blood
stream
either
directly
or
via
the
portal
vein.
Cyanogen
chloride
that
is
absorbed
via
the
portal
vein
would
be
expected
to
undergo
rapid
glutathione­
mediated
reduction
to
cyanide
and
HCl
in
the
blood
and
in
the
liver
during
first­
pass
metabolism,
in
light
of
the
rapid
reactivity
of
cyanogen
chloride
in
blood,
and
the
high
glutathione
concentration
in
the
liver.
Cyanide
produced
in
the
portal
vein
or
the
liver
would
be
expected
to
be
metabolized
to
thiocyanate
via
the
rhodanese
enzyme
in
the
liver.
Cyanogen
chloride
absorbed
directly
into
the
blood
stream
would
also
be
expected
to
react
rapidly
in
the
blood
to
form
cyanide
and
HCl.
Cyanide
formed
in
the
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
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Potential
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33
Final
draft
blood
following
intestinal
absorption
of
cyanogen
chloride
would
be
metabolized
by
rhodanese
found
elsewhere
in
the
body,
or
by
mercaptopyruvate
sulfur
transferase,
which
is
concentrated
in
blood
cells.

Based
on
this
hypothesized
toxicokinetic
pathway,
the
primary
metabolites
of
cyanogen
chloride
would
be
cyanide,
HCl,
and
thiocyanate,
and
consideration
of
the
toxic
effects
of
cyanogen
chloride
should
focus
on
these
compounds.
In
addition
to
effects
attributable
to
cyanide,
HCl,
and
thiocyanate,
primary
irritation
to
the
gastrointestinal
tract
might
also
occur,

based
on
the
reports
of
irritation
following
exposure
to
cyanogen
chloride
vapor
(
Flury
and
Zernick,
1931;
Prentiss,
1937;
Michigan
Department
of
Public
Health,
1977;
Reed,
1920;

Aldridge
and
Evans,
1946;
Haymaker
et
al.,
1952).
Insufficient
data
are
available
to
quantify
the
oral
doses
that
would
cause
this
effect,
but
it
is
reasonable
to
expect
that
this
irritation
would
occur
at
drinking
water
concentrations
above
those
that
could
result
in
systemic
effects.
The
hypothesized
metabolic
pathway
would
suggest
that
HCl
production
might
result
in
transient
pH
decreases,
particularly
in
the
portal
vein
and
liver
(
with
an
associated
potential
for
liver
damage),

depending
on
the
size
of
the
bolus
dose
and
the
buffering
capacity
of
these
tissues.
However,
no
effect
on
liver
weight,
liver
histopathology,
or
serum
enzymes
indicative
of
liver
damage
was
observed
in
rats
and
rabbits
at
drinking
water
doses
of
36
and
25
mg
HCl/
kg­
day,
respectively
(
Tober­
Meyer
et
al.,
1981);
there
was
also
no
acidosis
at
this
dose.
If
HCl
and
cyanogen
chloride
are
absorbed
to
a
similar
degree
and
at
similar
rates,
these
results
would
suggest
that
cyanogen
chloride
at
similar
molar
doses
would
also
not
result
in
liver
effects
from
HCl
production.

However,
uncertainties
remain
as
to
how
differences
between
HCl
and
cyanogen
chloride
in
the
rate
or
degree
of
absorption
from
the
stomach
and/
or
intestine
would
affect
the
peak
hydrogen
ion
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
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Potential
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34
Final
draft
concentration
in
the
liver.
Liver
effects
were
not
investigated
in
the
feed
studies
that
produced
acidosis
at
higher
doses
of
HCl
(
Throssell
et
al.,
1995;
1996),
but
liver
damage
does
not
appear
to
be
one
of
the
effects
associated
with
acidosis
(
Bookallil,
2001),
suggesting
that
the
liver
is
resistant
to
moderate
changes
in
blood
pH.
Liver
effects
were
not
reported
in
the
few
studies
on
cyanogen
chloride
inhalation
or
injection
(
Flury
and
Zernick,
1931;
Prentiss,
1937;
Michigan
Department
of
Public
Health,
1977;
Reed,
1920;
Aldridge
and
Evans,
1946;
Haymaker
et
al.,

1952),
but
the
liver
was
not
evaluated,
and
the
potential
liver
effects
from
cyanogen
chloride
ingestion
would
be
related
to
first­
pass
metabolism,
which
would
not
apply
to
these
inhalation
and
injection
studies.
Overall,
the
data
suggest
that
the
HCl
produced
from
metabolism
of
cyanogen
chloride
would
not
be
sufficient
to
damage
the
liver,
but
there
are
several
uncertainties
in
the
data.

A
second
possible
effect
of
HCl
production
from
cyanogen
chloride
might
be
systemic
metabolic
acidosis.
Due
to
the
complexities
of
determining
how
the
body
adjusts
to
increased
acid
in
the
system,
the
potential
for
acidosis
resulting
from
cyanogen
chloride
ingestion
is
discussed
below
in
the
context
of
specific
potential
exposure
levels.

Because
the
only
identified
metabolites
of
cyanogen
chloride
are
cyanide
and
thiocyanate,

consideration
of
appropriate
surrogates
was
limited
to
these
two
chemicals.
The
calculated
RfD
for
cyanide
is
0.005
mg/
kg­
day
based
on
the
NOAEL,
or
0.0008
mg/
kg­
day
based
on
the
BMDL.

For
thiocyanate,
the
calculated
RfD
is
0.001
mg/
kg­
day.
The
corresponding
values
for
the
cyanogen
chloride
RfD
that
would
be
calculated
from
the
cyanide
NOAEL­
based
RfD,
the
cyanide
BMDL­
based
RfD,
and
the
thiocyanate
RfD,
after
adjusting
for
the
molecular
weights
of
cyanogen
chloride
and
the
metabolites,
are
0.01,
0.002,
and
0.001
mg/
kg­
day,
respectively.
This
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
35
Final
draft
adjustment
is
based
on
a
determination
of
the
amount
of
the
metabolite
that
could
be
produced
from
a
given
dose
of
cyanogen
chloride,
assuming
that
all
of
the
cyanogen
chloride
is
converted
to
that
metabolite.

Cyanide
has
been
selected
as
the
surrogate
for
cyanogen
chloride
based
on
the
fact
that
it
is
a
known
metabolite
that
represents
a
major
portion
of
the
cyanogen
chloride
dose.
Although
the
lowest
cyanogen
chloride
RfD
would
have
resulted
if
thiocyanate
were
used
as
a
surrogate,

the
toxicokinetic
data
on
thiocyanate
and
cyanogen
chloride
are
insufficient
to
use
thiocyanate
as
the
surrogate.
There
are
no
quantitative
data
on
thiocyanate
production
from
cyanogen
chloride.

If
thiocyanate
were
to
be
used
as
a
surrogate,
quantitative
metabolic
data
(
including
mass­
balance
data)
on
the
degree
of
cyanogen
chloride
conversion
to
thiocyanate
at
environmentally­
relevant
doses
would
be
needed.
Therefore,
the
uncertainty
is
greater
when
thiocyanate
(
a
more
downstream
metabolite)
is
used
as
a
surrogate
instead
of
cyanide.
Given
that
a
significant
portion
of
an
ingested
cyanide
dose
would
react
or
be
excreted
prior
to
metabolism
to
thiocyanate,
the
derivation
of
a
thiocyanate
RfD
that
is
lower
than
the
cyanide
RfD
is
not
inconsistent
with
the
fact
that
cyanide
is
metabolized
to
thiocyanate.
Similarly,
a
portion
of
an
ingested
cyanogen
chloride
dose
may
be
converted
to
products
other
than
cyanide
or
thiocyanate.

If
the
cyanogen
chloride
RfD
were
calculated
using
thiocyanate
as
the
surrogate,
the
RfD
would
be
approximately
1/
10
the
RfD
calculated
using
cyanide
as
a
surrogate
(
when
the
RfD
is
based
on
the
NOAEL)
or
½
the
RfD
calculated
using
cyanide
as
a
surrogate
(
when
the
RfD
is
calculated
based
on
the
BMDL).
Therefore,
using
the
NOAEL­
based
cyanide
RfD,
cyanide
would
be
sufficiently
health­
protective
as
a
surrogate
unless
direct
conversion
to
thiocyanate
were
Drinking
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36
Final
draft
more
than
1/
10
of
the
cyanogen
chloride
dose.
Similarly,
using
the
BMDL­
based
cyanide
RfD,

cyanide
would
be
sufficiently
health­
protective
as
a
surrogate
unless
direct
conversion
to
thiocyanate
were
more
than
1/
2
of
the
cyanogen
chloride
dose.
Given
that
conversion
of
cyanogen
chloride
to
cyanide
is
dose­
related,
and
80%
of
cyanogen
chloride
is
converted
to
cyanide
at
low
doses,
it
is
likely
that
<
20%
of
cyanogen
chloride
is
converted
directly
to
thiocyanate
at
environmentally­
relevant
doses,
and
plausible
that
this
direct
conversion
accounts
for
<
10%
of
the
cyanogen
chloride
dose.
In
light
of
the
greater
uncertainty
involved
in
using
thiocyanate
as
a
surrogate,
and
given
that
the
factor
of
2
difference
between
the
20%
and
10%

conversion
is
well
within
the
uncertainty
of
the
method,
cyanide
is
preferred
as
the
surrogate
for
derivation
of
the
cyanogen
chloride
RfD.

Therefore,
the
quantitative
assessments
developed
for
cyanide
are
applied
to
cyanogen
chloride
by
adjusting
the
value
by
the
ratio
of
the
molecular
weights
of
cyanogen
chloride
and
cyanide,
61.5:
26,
(
approximately
2.36).
Confidence
in
the
use
of
cyanide
as
the
surrogate
RfD
is
reduced
because
the
critical
effect
for
cyanide,
male
reproductive
toxicity,
was
not
evaluated
in
the
limited
studies
on
cyanogen
chloride.
However,
it
is
plausible
that
male
reproductive
toxicity
could
be
observed
following
long­
term
exposure
to
cyanogen
chloride,
since
cyanide
is
a
major
metabolite.
It
should
be
noted
that
there
is
considerable
uncertainty
in
using
cyanide
as
a
surrogate
for
cyanogen
chloride,
due
to
the
incompleteness
of
the
data
on
cyanogen
chloride
toxicity
and
metabolism.
Indeed,
the
database
for
cyanogen
chloride
is
far
less
complete
than
the
minimal
database
required
for
deriving
even
a
"
low
confidence"
RfD.
Drinking
Water
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Document
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Cyanogen
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37
Final
draft
B.
3.1
One­
day
Health
Advisory
for
Cyanogen
Chloride
No
suitable
studies
are
available.
In
the
absence
of
adequate
data,
the
Ten­
day
HA
value
is
recommended
as
a
conservative
estimate
of
an
appropriate
One­
day
HA
value.

B.
3.2
Ten­
day
Health
Advisory
for
Cyanogen
Chloride
The
Ten­
day
Health
Advisory
for
cyanide
is
based
on
a
LOAEL
of
14
mg
CN/
kg­
day
for
increased
liver
weight
in
rats
that
ingested
cyanide
in
drinking
water
for
21
days
(
Palmer
and
Olson,
1979);
a
NOAEL
in
feed
in
the
same
study
was
not
used,
due
to
uncertainties
in
the
analysis
of
the
cyanide
in
feed.
Similarly,
a
BMDL
of
2.1
mg
CN/
kg­
day
identified
for
decreased
plasma
TSH
and
increased
thyroid
weight
in
a
rat
dietary
study
(
Kreutler
et
al.,
1978)
was
not
used
for
the
development
of
the
Health
Advisory,
due
to
the
large
degree
of
extrapolation
below
the
data.
An
uncertainty
factor
of
1000
was
used,
based
on
factors
of
10
each
for
interspecies
extrapolation,
protective
of
sensitive
subpopulations,
and
extrapolation
from
a
LOAEL
to
a
NOAEL.
This
value
is
used
for
cyanogen
chloride,
after
adjustment
for
the
relative
molecular
weights
of
cyanide
and
cyanogen
chloride.
(
See
Table
VIII­
7.)

Ten­
day
HA
for
child
=
(
0.14
mg/
L)
*
(
61.5/
26)
=
0.33
mg/
L
(
rounded
to
0.3
mg/
L)

where:

0.14
mg/
L
=
child
Ten­
day
HA
for
cyanide
(
prior
to
rounding)

61.5/
26
=
the
molecular­
weight
ratio
of
cyanogen
chloride
to
cyanide
Drinking
Water
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Document
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EPA/
OW/
OST/
HECD
VIII­
38
Final
draft
To
determine
whether
systemic
acidosis
resulting
from
HCl
production
may
be
a
concern
at
this
exposure
level,
the
number
of
mmoles
of
HCl
that
could
be
produced
from
this
dose
of
cyanogen
chloride
is
determined:

Dose
=
(
0.3
mg/
L)
(
1
L/
day)
=
0.0008
mmol
HCl/
kg­
day
(
10
kg)
(
36.5)

where:

0.3
mg/
kg­
day
=
Ten­
day
Health
Advisory
10
kg
=
assumed
body
weight
of
a
child.
1
L/
day
=
assumed
daily
water
consumption
by
a
10­
kg
child.
36.5
=
molecular
weight
of
HCl
For
comparison,
the
average
net
(
systemic)
acid
production
for
adults
is
about
60
milliequivalents
(
60
mmol)
per
day,
corresponding
to
approximately
0.9
mmol/
kg­
day
(
assuming
a
body
weight
of
70
kg).
Thus,
the
amount
of
HCl
produced
from
cyanogen
chloride
at
a
drinking
water
concentration
corresponding
to
the
Ten­
day
Health
Advisory
is
less
than
1/
1000
the
daily
acid
production,
and
would
be
toxicologically
inconsequential.
This
means
that
systemic
acidosis
would
not
be
a
concern.

B.
3.3
Longer­
term
Health
Advisory
for
Cyanogen
Chloride
The
Longer­
term
Health
Advisory
for
cyanide
is
based
on
the
finding
of
male
reproductive
effects
in
rats
at
a
NOAEL
of
4.5
mg/
kg­
day
(
NTP,
1993).
An
uncertainty
factor
of
1000
was
used,
based
on
factors
of
10
each
for
interspecies
extrapolation,
protection
of
sensitive
subpopulations,
and
for
an
incomplete
database
(
including
the
lack
of
a
2­
generation
reproduction
study,
and
lack
of
developmental
toxicity
studies,
particularly
the
absence
of
information
on
neurodevelopmental
effects).
An
alternative
Longer­
term
Health
Advisory
for
cyanide
was
developed
based
on
a
BMDL
of
0.79
mg/
kg­
day
for
decreased
epididymis
weight
in
the
NTP
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
39
Final
draft
(
1993)
study,
using
the
same
uncertainty
factors.
The
Longer­
term
Health
Advisory
based
on
the
NOAEL,
as
well
as
the
one
based
on
the
BMDL,
were
used
for
cyanogen
chloride,
after
adjusting
for
the
relative
molecular
weights
of
cyanide
and
cyanogen
chloride.
An
additional
modifying
factor
to
account
for
production
of
HCl
is
not
needed,
because
the
critical
effect
occurs
in
the
male
reproductive
tract,
while
the
liver
is
the
site
where
HCl
formed
from
cyanogen
chloride
would
be
formed
and
would
be
the
target
of
most
concern
for
HCl
exposure.

Derivation
of
the
Longer­
term
HA
based
on
the
study
NOAEL
Longer­
term
HA
for
child
=
(
0.045
mg/
L)
(
61.5/
26)
=
0.106
mg/
L
(
rounded
to
0.1
mg/
L)
where:
0.045
mg/
L
=
child
Longer­
term
HA
for
cyanide
(
prior
to
rounding)
61.5/
26
=
the
molecular­
weight
ratio
of
cyanogen
chloride
to
cyanide
Longer­
term
HA
for
adult
=
(
0.158
mg/
L)
(
61.5/
26)
=
0.37
mg/
L
(
rounded
to
0.4
mg/
L)
where:
0.158
mg/
L
=
adult
Longer­
term
HA
for
cyanide
(
prior
to
rounding)
61.5/
26
=
the
molecular
weight
ratio
of
cyanogen
chloride
to
cyanide
Derivation
of
the
Longer­
term
HA
based
on
the
study
BMDL
Longer­
term
HA
for
child
=
(
0.0079
mg/
L)
(
61.5/
26)
=
0.019
mg/
L
(
rounded
to
0.02
mg/
L)
where:
0.0079
mg/
L
=
child
Longer­
term
HA
for
cyanide
(
prior
to
rounding)
61.5/
26
=
the
molecular­
weight
ratio
of
cyanogen
chloride
to
cyanide
Longer­
term
HA
for
adult
=
(
0.028
mg/
L)
(
61.5/
26)
=
0.066
mg/
L
(
rounded
to
0.07
mg/
L)
where:
0.028
mg/
L
=
adult
Longer­
term
HA
for
cyanide
(
prior
to
rounding)
61.5/
26
=
the
molecular
weight
ratio
of
cyanogen
chloride
to
cyanide
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
40
Final
draft
B.
3.4
Reference
Dose,
Drinking
Water
Equivalent
Level,
and
Lifetime
Health
Advisory
for
Cyanogen
Chloride
The
RfD
for
cyanide
is
based
on
the
finding
of
male
reproductive
effects
(
decreased
epididymis
and
testes
weight,
decreased
spermatid
counts)
in
rats
at
a
NOAEL
of
4.5
mg/
kg­
day
(
NTP,
1993).
An
uncertainty
factor
of
1000
was
used,
based
on
factors
of
10
each
for
interspecies
extrapolation
and
protection
of
sensitive
subpopulations,
and
a
factor
of
10
for
subchronic
to
chronic
extrapolation
and
database
considerations
combined.
An
alternative
RfD
for
cyanide
was
developed
based
on
a
BMDL
of
0.79
mg/
kg­
day
for
decreased
epididymis
weight
in
the
NTP
(
1993)
study,
using
the
same
uncertainty
factors.
The
RfD
based
on
the
NOAEL,
as
well
as
the
one
based
on
the
BMDL,
were
used
for
cyanogen
chloride,
after
adjusting
for
the
relative
molecular
weights
of
cyanide
and
cyanogen
chloride.
An
additional
modifying
factor
to
account
for
production
of
HCl
is
not
needed,
because
the
critical
effect
occurs
in
the
male
reproductive
tract,
while
the
liver
is
the
site
where
HCl
formed
from
cyanogen
chloride
would
be
formed
and
would
be
the
target
of
most
concern
for
HCl
exposure.

Derivation
of
the
Lifetime
HA
based
on
the
study
NOAEL
Step
1:
Determination
of
RfD
for
Cyanogen
chloride.

RfD
=
(
0.0045
mg/
kg­
day)
(
61.5/
26)
=
0.011
mg/
kg­
day
(
rounded
to
0.01
mg/
kg­
day)

where:
0.0045
mg/
kg­
day
=
the
RfD
for
cyanide
(
before
rounding)
61.5/
26
=
the
molecular­
weight
ratio
of
cyanogen
chloride
to
cyanide
Step
2:
Determination
of
a
Drinking
Water
Equivalent
Level
(
DWEL)
for
Cyanogen
chloride
DWEL
=
(
0.011
mg/
kg­
day)
(
70
kg)
=
0.39
mg/
L
(
rounded
to
0.4
mg/
L)
(
2
L/
day)
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
41
Final
draft
where:

0.011
mg/
kg­
day
=
RfD
(
prior
to
rounding)
70
kg
=
assumed
body
weight
of
an
adult
2
L/
day
=
assumed
drinking
water
consumption
of
a
70­
kg
adult
Step
3:
Determination
of
Lifetime
HA
for
cyanogen
chloride
Lifetime
HA
=
(
0.39
mg/
L)
(
20%)
=
0.078
mg/
L
(
rounded
to
80

g/
L)

where:
0.39
mg/
L
=
DWEL
(
before
rounding)
20%
=
assumed
relative
source
contribution
from
water
Derivation
of
the
Lifetime
HA
based
on
the
study
BMDL
Step
1:
Determination
of
RfD
for
Cyanogen
chloride.

RfD
=
(
0.00079
mg/
kg­
day)
(
61.5/
26)
=
0.0019
mg/
kg­
day
(
rounded
to
0.002
mg/
kg­
day)

where:
0.00079
mg/
kg­
day
=
the
RfD
for
cyanide
(
before
rounding)
61.5/
26
=
the
molecular­
weight
ratio
of
cyanogen
chloride
to
cyanide
Step
2:
Determination
of
a
Drinking
Water
Equivalent
Level
(
DWEL)
for
Cyanogen
chloride
DWEL
=
(
0.0019
mg/
kg­
day)
(
70
kg)
=
0.067
mg/
L
(
rounded
to
0.07
mg/
L)
(
2
L/
day)

where:

0.0019
mg/
kg­
day
=
RfD
(
prior
to
rounding)
70
kg
=
assumed
body
weight
of
an
adult
2
L/
day
=
assumed
drinking
water
consumption
of
a
70­
kg
adult
Step
3:
Determination
of
Lifetime
HA
for
cyanogen
chloride
Lifetime
HA
=
(
0.067
mg/
L)
(
20%)
=
0.013
mg/
L
(
rounded
to
10

g/
L)

where:
0.067
mg/
L
=
DWEL
(
before
rounding)
20%
=
assumed
relative
source
contribution
from
water
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
42
Final
draft
C.
Carcinogenic
Effects
Since
there
are
no
cancer
bioassays
of
cyanogen
chloride,
the
carcinogenicity
assessment
for
this
chemical
is
also
conducted
by
evaluation
of
its
known
and
putative
metabolites.
There
are
no
cancer
bioassays
for
cyanide
or
cyanate.
There
are
no
well­
conducted
standard
cancer
bioassays
of
thiocyanate.
Two
oral
carcinogenicity
studies
of
thiocyanate
in
male
Fischer
rats
were
conducted
in
the
same
laboratory
(
Lijinsky
and
Reuber,
1982;
Lijinsky
and
Kovatch,
1989).

The
only
effect
in
the
first
study
was
an
increase
in
liver
tumors,
and
this
was
not
confirmed
at
the
higher
dose
tested
in
the
second
study,
although
a
non­
statistically
significant
increase
in
thyroid
tumors
was
observed.
However,
these
studies
suffer
from
a
number
of
limitations.
Only
a
single
dose
was
tested
in
each
study,
only
20
rats/
group
were
tested,
and
the
assay
continued
until
all
of
the
animals
died
(
increasing
the
background
incidence
of
age­
related
tumors,
and
thus
decreasing
the
sensitivity
of
the
bioassay),
and
results
were
not
fully
reported.

As
described
in
Appendix
D,
one
standard
cancer
bioassay
on
cyanamide
was
located.

NCI
(
1979)
evaluated
the
carcinogenicity
of
calcium
cyanamide
in
drinking
water
to
F344
rats
and
B6C3F1
mice.
No
tumors
related
to
treatment
were
observed
in
rats.
Hemangiosarcomas
were
observed
in
male
mice
with
a
significant
dose­
related
trend.
However,
the
tumor
incidence
was
not
statistically
significant
in
either
dose
group
when
compared
with
controls.
Malignant
lymphomas
were
observed
in
female
mice.
Statistical
analysis
indicated
a
significant
dose­
related
trend
for
these
tumors
and
a
significant
increase
in
incidence
in
the
high­
dose
group
compared
with
controls.
However,
NCI
(
1979)
reported
that
the
historical
control
for
this
tumor
type
in
B6C3F1
mice
was
21%;
therefore,
the
incidence
of
5%
in
the
matched
control
group
may
have
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
43
Final
draft
been
abnormally
low.
NCI
(
1979)
concluded
that
neither
of
these
tumors
was
clearly
related
to
administration
of
calcium
cyanamide.

No
data
are
available
on
the
genotoxicity
of
cyanogen
chloride,
although
a
QSTR
analysis
predicted
that
cyanogen
chloride
was
negative
for
carcinogenicity
in
rats
and
mice
of
both
sexes
(
Moudgal
et
al.,
2000).
Although
the
data
for
some
known
or
potential
metabolites
of
cyanogen
chloride
are
limited,
the
available
data
suggest,
overall,
that
none
of
the
metabolites
is
genotoxic.

Overall,
cyanide
has
been
negative
in
bacterial­
mutagenicity
assays
(
De
Flora,
1981;
De
Flora
et
al.,
1984;
NTP,
1993)
and
assays
of
DNA
damage
and
repair
(
De
Flora
et
al.,
1984;
Painter
and
Howard,
1982),
although
a
positive
result
was
obtained
in
one
strain
of
Salmonella
with
and
without
S9
activation
(
Kushi
et
al.,
1983).
No
assays
of
mammalian
gene
mutation
or
chromosome
aberration
are
available.
Genotoxicity
data
on
thiocyanate
are
limited
to
marginal
results
in
a
S.
typhimurium­
mutagenicity
assay
(
Kier,
1988,
as
reported
by
Rosenkranz
and
Klopman,
1990).
Potassium
thiocyanate
does
not
have
any
structural
alerts
for
genotoxicity
(
Rosenkranz
and
Klopman,
1990).
Cyanate
was
negative
in
the
only
study
located,
a
mammalian
gene­
mutation
assay
(
Melzer
et
al.,
1983).
Except
for
a
positive
result
in
Salmonella
strain
TA
1535
(
Zeiger,
1987),
cyanamide
was
negative
in
bacterial
gene­
mutation
assays
(
Loveless
et
al.,

1954;
Zeiger,
1987),
a
Drosophila
sex­
linked
recessive
assay
(
Yoon
et
al.,
1985),
an
assay
of
DNA
damage
in
mammalian
cells
(
Sina
et
al.,
1983),
and
a
micronucleus
assay
in
mice
(
Menargues
et
al.,
1984).

Based
on
these
considerations,
cyanogen
chloride,
cyanide,
thiocyanate,
cyanate,
and
cyanamide
are
all
classified
as
Group
D,
Not
Classifiable
as
to
Human
Carcinogenicity,
using
the
Drinking
Water
Criteria
Document
for
Cyanogen
Chloride
and
Potential
Metabolites
EPA/
OW/
OST/
HECD
VIII­
44
Final
draft
U.
S.
EPA
(
1986)
guidelines.
Using
the
U.
S.
EPA
(
1999)
Draft
Guidelines
for
Carcinogen
Risk
Assessment,
the
data
are
inadequate
for
an
assessment
of
the
human
carcinogenic
potential
of
these
compounds.

D.
Characterization
of
Uncertainties
and
Data
Gaps
Very
little
toxicity
information
exists
for
cyanogen
chloride,
and
the
existing
data
on
cyanogen
chloride
were
developed
primarily
prior
to
the
advent
of
modern
methods
of
toxicology.

This
assessment
assumed
that,
based
on
the
available
data
on
cyanogen
chloride
metabolism,

structural
similarity,
and
qualitative
similarity
of
observed
endpoints,
it
is
appropriate
to
use
cyanide
as
a
surrogate
for
cyanogen
chloride
toxicity.
However,
that
assumption
has
not
been
tested
directly.
Ideally,
sufficient
data
from
long­
term
studies
would
be
obtained
so
that
the
toxicity
values
can
be
derived
from
data
on
cyanogen
chloride,
rather
than
a
surrogate.
Failing
that,
data
from
short­
term
toxicity
studies
can
be
used
to
identify
targets
of
toxicity
and
to
test
the
hypothesis
that
cyanide
is
an
appropriate
surrogate.
High­
quality,
in
vivo
metabolism
studies
(
or
at
least
in
vitro
metabolism
studies
using
cell
extracts)
would
also
be
useful
to
determine
which
of
the
predicted
metabolites
actually
are
formed,
and
to
quantitate
their
production.
The
NTP
has
recently
selected
cyanogen
chloride
for
testing;
such
testing
would
significantly
reduce
the
uncertainties
in
this
assessment.

There
are
also
uncertainties
in
the
assessments
for
some
of
the
metabolites,
as
discussed
in
the
context
of
the
database
uncertainty
factor
for
each
RfD,
and
as
discussed
in
Chapter
9.

Confidence
in
the
calculated
health
advisories
is
also
discussed
in
Chapter
9.
