i
United
States
Prevention,
Pesticides
Environmental
Protection
and
Toxic
Substances
February
8,
2006
Agency
(
7508C)

Assessment
of
Lindane
and
Other
Hexachlorocyclohexane
Isomers
February
8,
2006
ii
TABLE
OF
CONTENTS
Lindane
Assessment
Team..........................................................................................................
iii
Glossary
of
Terms
and
Abbreviations
...........................................................................................
iv
I.
Introduction.........................................................................................................................
1
A.
History
and
Regulatory
Status
of
Lindane
in
the
U.
S.
.....................................................
1
B.
Global
Status
of
HCH......................................................................................................
3
II.
Properties
and
Production
of
Hexachlorocyclohexanes
(
HCH).........................................
9
A.
Chemical
Identity
............................................................................................................
9
B.
Manufacturing
Process
..................................................................................................
11
C.
Global
Use
and
Production
............................................................................................
13
III.
HCH
Environmental
Fate
and
Effects
Risk
Assessment
.....................................................
15
A.
Environmental
Fate
and
Transport....................................................................................
15
B.
Ecological
Effects..........................................................................................................
22
IV.
HCH
Human
Health
Risk
Assessment..............................................................................
26
A.
Gamma­
HCH
Isomer
­
Lindane
.......................................................................................
26
B.
Alpha­
and
Beta­
HCH
Isomers
......................................................................................
27
V.
Additional
Concerns
and
Information
Request...................................................................
50
VI.
Bibliography
.....................................................................................................................
52
Appendix
A.......................................................................................................................
66
iii
Lindane
and
Other
Hexachlorocyclohexane
Isomers
Assessment
Team
Biological
and
Economic
Analysis
Assessment
Dave
Brassard
Environmental
Fate
and
Effects
Risk
Assessment
Nicholas
Federoff
Faruque
Khan
Field
and
External
Affairs
Division
Janice
Jensen
Health
Effects
Risk
Assessment
Rebecca
Daiss
Thurston
Morton
Registration
Support
George
LaRocca
Reregistration
Support
Mark
T.
Howard
Kimberly
Nesci
Office
of
General
Counsel
Gautam
Srinivasan
iv
Glossary
of
Terms
and
Abbreviations
a.
i.
Active
Ingredient
aPAD
Acute
Population
Adjusted
Dose
BCF
Bioconcentration
Factor
CDC
Centers
for
Disease
Control
CDPR
California
Department
of
Pesticide
Regulation
CFR
Code
of
Federal
Regulations
cPAD
Chronic
Population
Adjusted
Dose
CSFII
USDA
Continuing
Surveys
for
Food
Intake
by
Individuals
CWS
Community
Water
System
DCI
Data
Call­
In
DEEM
Dietary
Exposure
Evaluation
Model
EDSP
Endocrine
Disruptor
Screening
Program
EDSTAC
Endocrine
Disruptor
Screening
and
Testing
Advisory
Committee
EEC
Estimated
Environmental
Concentration.
EP
End­
Use
Product
EPA
U.
S.
Environmental
Protection
Agency
EXAMS
Tier
II
Surface
Water
Computer
Model
FDA
Food
and
Drug
Administration
FFDCA
Federal
Food,
Drug,
and
Cosmetic
Act
FIFRA
Federal
Insecticide,
Fungicide,
and
Rodenticide
Act
FOB
Functional
Observation
Battery
FQPA
Food
Quality
Protection
Act
FR
Federal
Register
GPS
Global
Positioning
System
HIARC
Hazard
Identification
Assessment
Review
Committee
IDFS
Incident
Data
System
RED
Reregistration
Eligibility
Decision
LADD
Lifetime
Average
Daily
Dose
LC50
Median
Lethal
Concentration.
Statistically
derived
concentration
of
a
substance
expected
to
cause
death
in
50%
of
test
animals,
usually
expressed
as
the
weight
of
substance
per
weight
or
volume
of
water,
air
or
feed,
e.
g.,
mg/
l,
mg/
kg
or
ppm.
LD50
Median
Lethal
Dose.
Statistically
derived
single
dose
causing
death
in
50%
of
the
test
animals
when
administered
by
the
route
indicated
(
oral,
dermal,
inhalation),
expressed
as
a
weight
of
substance
per
unit
weight
of
animal,
e.
g.,
mg/
kg.
LOAEC
Lowest
Observed
Adverse
Effect
Concentration
LOAEL
Lowest
Observed
Adverse
Effect
Level
LOC
Level
of
Concern
LOEC
Lowest
Observed
Effect
Concentration
mg/
kg/
day
Milligram
Per
Kilogram
Per
Day
MOE
Margin
of
Exposure
MP
Manufacturing­
Use
Product
MRID
Master
Record
Identification
(
number).
EPA's
system
of
recording
and
tracking
studies
submitted.
MRL
Maximum
Residue
Level
N/
A
Not
Applicable
NARAP
North
American
Regional
Action
Plan
NASS
National
Agricultural
Statistical
Service
NAWQA
USGS
National
Water
Quality
Assessment
NMFS
National
Marine
Fisheries
Service
NOAEC
No
Observed
Adverse
Effect
Concentration
NOAEL
No
Observed
Adverse
Effect
Level
NPIC
National
Pesticide
Information
Center
OP
Organophosphorus
OPP
EPA
Office
of
Pesticide
Programs
v
ORETF
Outdoor
Residential
Exposure
Task
Force
PAD
Population
Adjusted
Dose
PCA
Percent
Crop
Area
PDCI
Product
Specific
Data
Call­
In
PDP
USDA
Pesticide
Data
Program
ppb
Parts
Per
Billion
PPE
Personal
Protective
Equipment
PRZM
Pesticide
Root
Zone
Model
RBC
Red
Blood
Cell
RED
Reregistration
Eligibility
Decision
REI
Restricted
Entry
Interval
RfD
Reference
Dose
RPA
Reasonable
and
Prudent
Alternatives
RPM
Reasonable
and
Prudent
Measures
RQ
Risk
Quotient
RTU
(
Ready­
to­
use)
RUP
Restricted
Use
Pesticide
SCI­
GROW
Tier
I
Ground
Water
Computer
Model
SF
Safety
Factor
SLN
Special
Local
Need
(
Registrations
Under
Section
24(
c)
of
FIFRA)
STORET
Storage
and
Retrieval
TEP
Typical
End­
Use
Product
TGAI
Technical
Grade
Active
Ingredient
TRAC
Tolerance
Reassessment
Advisory
Committee
UF
Uncertainty
Factor
USDA
United
States
Department
of
Agriculture
USFWS
United
States
Fish
and
Wildlife
Service
USGS
United
States
Geological
Survey
1
I.
Introduction
A.
History
and
Regulatory
Status
of
Lindane
in
the
U.
S.

Lindane
was
first
registered
as
a
pesticide
in
the
United
States
in
the
1940s
for
veterinary
uses.
It
was,
at
that
time,
known
as
benzene
hexachloride
(
BHC).
Over
time,
lindane
was
registered
for
use
on
a
wide
variety
of
food
crops,
ornamental
plants,
livestock
and
homeowner
and
other
sites.
In
1977,
EPA
initiated
a
Rebuttable
Presumption
Against
Registration
(
RPAR)
review
of
lindane,
now
called
a
Special
Review.
As
a
part
of
the
Special
Review,
EPA
published
Position
Documents
from
1977
through
1983
resulting
in
the
cancellation
of
certain
uses
of
lindane.
In
1979,
technical
hexachlorocyclohexane
(
HCH)
was
no
longer
permitted
to
be
used
as
a
pesticide
in
the
US
and
the
lindane
sold
was
gamma
HCH.

EPA
issued
a
Registration
Standard
for
Lindane
in
September
1985
that
included
a
requirement
for
the
submission
of
additional
data
to
support
lindane
registration
and
to
address
exposure
concerns.
In
1998
and
1999,
lindane
registrants
voluntarily
cancelled
all
registered
uses
of
lindane
except
for
seed
treatment
uses
on
19
agricultural
crops
and
a
dog
mange
treatment.
In
2001
and
2002,
registrants
voluntarily
cancelled
the
lindane
mange
treatment
and
seed
treatment
use
on
13
crops
leaving
only
seed
treatment
use
on
six
agricultural
crops,
as
follows:
barley,
corn,
oats,
rye,
sorghum,
and
wheat.
Lindane
is
generally
used
to
control
wireworms
(
click
beetle
grubs)
and
flea
beetles.

In
addition
to
the
above
actions
on
the
registration
of
lindane,
in
1999,
the
US
nominated
lindane
for
the
preparation
of
a
North
American
Regional
Action
Plan
(
NARAP).

On
July
31,
2002,
EPA
completed
its
Reregistration
Eligibility
Decision
(
RED)
for
lindane
on
the
six
remaining
agricultural
uses
of
lindane,
seed
treatment
on
barley,
corn,
oats,
rye,
sorghum,
and
wheat.
The
RED
document
states
that
the
six
remaining
lindane
seed
treatment
uses
are
eligible
for
reregistration
provided
that
1)
registrants
make
required
label
changes;
2)
registrants
provide
required
data;
and
3)
the
Agency
is
able
to
establish
all
required
tolerances
for
lindane
residues
on
food.

To
address
the
first
condition
of
the
RED,
revised
lindane
end­
use
product
labels
were
provided
to
the
Agency,
and
the
Agency
has
reviewed
and
approved
these
labels
to
ensure
that
product
labeling
reflects
the
risk
mitigation
measures
stipulated
in
the
RED.

To
address
the
second
condition
of
the
RED,
registrants
have
since
provided
all
of
the
required
data.
The
registrants
have
submitted
the
required
product
and
residue
chemistry
data,
and
the
Agency
has
reviewed
these
data
and
found
them
to
be
acceptable.
The
registrants
have
also
submitted
an
outstanding
nature
of
the
residue
study,
also
known
as
a
plant
metabolism
study,
originally
required
in
the
1985
Registration
Standard
DCI
for
lindane,
and
these
data
are
currently
in
review.
As
stipulated
in
the
RED
document,
the
Agency
may
require
additional
residue
chemistry
studies
if
residues
of
2
concern
other
than
lindane
are
detected
in
the
nature
of
the
residue
study.
The
registrants
have
submitted
a
required
seed
leaching
study
that
has
been
reviewed
and
found
to
be
acceptable
by
the
Agency.
And
finally,
the
registrant
has
submitted
an
anaerobic
aquatic
metabolism
study
to
satisfy
an
anaerobic
soil
metabolism
data
requirement
originally
required
under
the
DCI
from
the
1985
Lindane
Reregistration
Standard.
(
EPA
now
typically
requests
anaerobic
aquatic
metabolism
studies
in
place
of
anaerobic
soil
metabolism
studies
because
an
anaerobic
soil
metabolism
study
is
not
needed
if
an
anaerobic
aquatic
metabolism
study
is
available;
therefore,
one
study
will
satisfy
both
data
requirements.)
The
Agency
is
currently
reviewing
this
study.

As
a
result
of
the
Agency's
continuing
review
of
lindane,
the
Agency
initiated
the
preparation
of
this
document.
This
document
presents
EPA's
revised
assessment
of
risks
related
to
the
continued
registration
of
the
insecticide
lindane,
also
known
as
gamma
HCH.

Consistent
with
comments
received
on
the
lindane
RED
(
Docket
Control
Number
OPP­
2002­
0202)
and
consistent
with
the
draft
North
American
Regional
Action
Plan
(
NARAP)
on
Lindane
and
Other
Hexachlorocyclohexane
Isomers,
the
Agency
is
considering
risks
resulting
from
human
and
environmental
exposures
to
other
HCH
isomers
of
environmental
significance
produced
as
by­
products
during
the
manufacture
of
lindane.

In
addition
to
its
agricultural
uses,
lindane
is
approved
by
the
US
Food
and
Drug
Administration
(
FDA)
for
lice
and
scabies
treatment
and
has
been
marketed
as
a
pharmaceutical
product
since
1951.
In
2003,
as
a
result
of
the
reassessment
of
lindane
risk
factors,
FDA
issued
a
Public
Health
Advisory
(
PHA)
on
the
safety
of
topical
lindane
products
and
took
action
to
increase
hazard
warnings
emphasizing
that
lindane
is
a
second­
line
treatment
and
to
reduce
the
maximum
package
size
to
minimize
the
possibility
of
overuse.
See
also
the
following
web
site
for
the
PHA
additional
information:
http://
www.
fda.
gov/
cder/
drug/
infopage/
lindane/
lindanePHA.
htm.

1.
Pesticidal
Uses
Currently,
the
only
registered
agricultural
uses
for
lindane
are
the
six
seed
treatment
uses
(
barley,
corn,
oats,
rye,
sorghum,
and
wheat),
as
stated
above.
Current
usage
of
lindane
in
the
US
is
less
than
150,000
lbs
of
active
ingredient
(
ai)
annually
to
about
9.7
million
acres.
The
percent
of
seeds
treated
is
low
(
7%
for
wheat
and
barley;
1%
for
oats
and
rye;
6%
for
corn;
and
1%
for
sorghum).

As
of
2002,
imidacloprid
and
thiamethoxam
were
the
primary
seed
treatment
alternatives
to
lindane
for
barley,
corn,
sorghum,
and
wheat.
These
alternatives
are
as
effective
as
lindane,
but
costlier
to
use.
In
addition,
since
2002,
additional
seed
treatment
alternatives
have
been
registered
on
corn
and
sorghum
(
clothianidin
on
corn
and
sorghum
and
thiamethoxam
on
corn).
Because
of
the
availability
of
these
alternatives,
growerlevel
impacts
if
lindane
was
no
longer
available
for
seed
treatment
use
on
barley,
corn,
sorghum,
and
wheat
would
be
minor.
For
wheat
and
barley,
the
estimated
increase
in
3
treatment
cost
would
be
$
0.36
to
$
1.71
per
acre
($
5
million
for
all
US
acreage,
or
about
0.06%
of
total
US
crop
value).
For
corn,
the
estimated
increase
in
treatment
cost
would
be
$
1.82
per
acre
($
8.7
million
dollars
for
all
US
acreage,
or
about
0.04
%
of
total
US
crop
value).
For
sorghum,
the
estimated
increase
in
treatment
cost
would
be
$
3.70
to
$
4.69
per
acre
(
about
$
386,000
for
all
US
acreage,
or
about
0.04
%
of
total
US
crop
value).

There
are
no
registered
alternatives
for
oats
and
rye;
therefore,
if
lindane
was
no
longer
available,
grower
level
impacts
would
be
major,
estimated
at
a
9%
yield
loss
(
about
$
0.77
per
acre,
resulting
in
$
310,000
or
0.2%
of
total
crop
value).
However,
an
application
is
pending
at
the
Agency
for
the
registration
of
imidacloprid
as
a
seed
treatment
on
oats
and
rye.
A
Federal
Register
notice
has
been
published
proposing
that
a
tolerance
be
established
for
these
uses
(
71
FR
4580).
Provided
that
imidacloprid
is
registered
for
use
on
oat
and
rye
seed,
grower­
level
impacts
of
the
loss
of
lindane
for
these
two
seed
treatments
would
also
be
minor,
with
an
increase
in
treatment
costs
estimated
at
$
0.36
to
$
1.71
per
acre
($
60,000
for
all
US
acreage,
or
about
0.04%
of
total
US
crop
value).

2.
Pharmaceutical
Uses
As
stated
above,
lindane
is
used
in
the
US
for
pediculosis,
lice
and
scabies
treatment.
Annual
use
of
lindane
as
a
pharmaceutical
to
treat
lice
and
scabies
in
the
US
is
less
than
one
metric
ton
(
or
<
1,000
kg).
Lindane
accounts
for
fewer
than
1
million
treatments
out
of
10
to
20
million
annual
cases
of
lice.
In
addition,
FDA
has
established
processes
for
facilitating
development
and
approving
the
use
of
botanicals
and
other
proposed
lice
and
scabies
treatments
for
pharmaceutical
purposes,
thereby
encouraging
the
use
of
lindane
alternatives.

B.
Global
Status
of
HCH
Based
on
information
collected
from
a
variety
of
sources,
lindane
is
banned
for
use
in
52
countries,
restricted
or
severely
restricted
in
33
countries,
not
registered
in
10
countries,
and
registered
in
17
countries.
A
summary
list
of
these
countries
is
included
as
Appendix
A
to
this
document
(
2005
draft
NARAP;
information
found
at
www.
cec.
org.)

1.
Regulatory
Status
in
Canada
Historically,
lindane
had
been
registered
in
Canada
for
a
wide
variety
of
applications.
Lindane
has
never
been
produced
in
Canada;
therefore,
Canada
has
imported
all
technical­
grade
lindane
from
foreign
companies.
Publication
of
Trade
Memorandum
T­
68
on
November
5,
1970,
signaled
an
end
to
the
use
of
lindane
on
a
range
of
fruit
and
vegetable
crops,
in
outdoor
foggers,
and
for
the
treatment
of
water
for
control
of
mosquitoes.
By
1999,
most
of
the
above­
ground
uses
of
lindane
in
Canada
were
discontinued.
Of
the
remaining
uses,
canola
seed
treatment
represented
greater
than
90%
of
lindane
use.
4
In
March
1999,
the
Health
Canada's
Pest
Management
Regulatory
Agency
(
PMRA)
announced
the
review
(
Special
Review)
of
the
remaining
seed
treatment
uses
of
lindane.
As
a
signatory
of
an
international
agreement
aiming
at
restricting
or
eliminating
Persistent
Organic
Pollutants
(
POPs)
such
as
lindane,
Canada
was
committed
to
conduct
a
reassessment
of
all
uses
of
lindane.

The
PMRA
completed
its
assessment
of
occupational
risk
in
October
2001.
The
use
of
lindane
was
phased
out
on
the
basis
of
unacceptable
risk
to
the
health
of
workers
exposed
to
lindane
during
seed
treatment
and
planting.
Crops
for
which
alternatives
were
already
available
were
allowed
to
be
used
until
October
1,
2002.
Crops
for
which
no
alternatives
were
yet
available
were
to
be
phased
out
(
last
use)
by
December
31,
2004.
All
registrants
of
lindane
seed
treatment
products,
except
one,
chose
to
voluntarily
discontinue
sales
of
their
products.

As
is
their
right
under
Section
23
of
the
Pest
Control
Product
Regulations,
the
registrant
that
did
not
voluntarily
agree
to
discontinue
sales
requested
a
hearing
by
an
independent
board
to
review
the
PMRA
decision
with
respect
to
its
lindane
products.
On
August
17,
2005,
the
Board
made
recommendations
to
the
Canadian
Minister
of
Health
to
have
the
PMRA
work
with
the
lindane
registrant
who
filed
the
Review
Board's
request
to
take
into
account
any
relevant
mitigation
measures
available
and
to
consider
the
possibility
of
a
mitigation
strategy
that
might
result
in
labels
and
uses
practices
acceptable
to
the
PMRA.
At
the
time
of
preparation
of
this
document,
the
PMRA
has
not
responded
in
public
to
the
Board's
report.

Lindane
is
approved
in
Canada
for
lice
and
scabies
treatment
as
a
non
prescription
"
behind
the
counter"
drug
with
5
commercial
products
containing
1%
lindane
in
solution,
currently
marketed
by
3
companies.
Lindane
has
been
registered
in
Canada
as
a
pharmaceutical
product
since
the
early
1960s,
and
this
use
is
the
only
current
allowable
use
of
lindane
in
Canada.
Lindane
products
have
been
classified
as
Schedule
2
products
by
the
National
Association
of
Pharmacy
Regulatory
Authorities
(
NAPRA),
which
means
that
"
professional
intervention
from
the
pharmacist
at
the
point
of
sale
and
possibly
referral
to
a
practitioner"
is
required.
The
product
is
available
only
from
a
pharmacist,
over­
the­
counter,
and
must
be
retained
within
an
area
of
the
pharmacy
where
there
is
no
public
access
and
no
opportunity
for
patient
self­
selection.
Provincial
pharmacist
associations
that
are
not
currently
members
of
NAPRA
(
Quebec
and
Ontario)
follow
similar
practices
and
guidelines.
Labelling
requirements
for
Canadian
lindane
pharmaceutical
products
are
available
on­
line:

Lindane
Lotion:
http://
www.
hc­
sc.
gc.
ca/
hpfb­
dgpsa/
tpd­
dpt/
Lindanel_
e.
html
Lindane
Shampoo:
http://
www.
hc­
sc.
gc.
ca/
hpfb­
dgpsa/
tpd­
dpt/
Lindanes_
e.
html
In
2003,
the
pharmaceutical
use
amounted
to
approximately
6
kg
of
lindane
per
year.
With
the
introduction
of
other
agents
such
as
permethrin,
the
use
of
lindane
has
5
declined
over
the
years.
This
current
use
of
6
kg
per
year
of
lindane
is
not
a
significant
source,
representing
0.005%
to
0.007%
of
the
North
American
total
use.

2.
Regulatory
Status
in
Mexico
(
Draft
NARAP,
2005)

In
2005,
Mexico
decided
to
eliminate
all
agricultural,
veterinary,
and
pharmaceutical
uses
of
lindane
through
a
prioritized,
phase­
out
approach.
Reasonable
timeframes
for
a
voluntary
phase
out
are
currently
being
negotiated
between
COFEPRIS
(
Federal
Commission
for
Sanitary
Risks
Protection,
Ministry
of
Health)
and
industry.
There
are
no
reported
exports
of
lindane
from
Mexico
to
other
countries
and
imports
of
the
active
ingredient
are
now
prohibited.
As
of
January
2005,
pollutant
release
and
transfer
register
(
PRTR)
reporting
is
mandatory
for
industry
in
Mexico.
Lindane
is
listed
in
Mexico's
PRTR
as
a
substance
for
voluntary
reporting
and
is
presently
being
considered
for
addition
to
the
mandatory
reporting
list.
In
2004,
Mexico
released
the
Mexican
National
Diagnostic
on
Lindane
(
See
http://
www.
ine.
gob.
mx/
dgicurg/
download/
Proyectos­
2003/
EL_
LINDANO_
EN_
MEXICO.
pdf)
to
support
activities
for
the
NARAP,
and
prepared
a
National
Implementation
Plan
for
Persistent
Organic
Pollutants
(
POP)
management
under
the
Stockholm
Convention.
In
2005,
Mexico
submitted
a
proposal
to
list
lindane
in
the
Annex
A
(
Elimination)
of
the
Stockholm
Convention.

Lindane
is
listed
in
the
Comisión
Intersecretarial
para
el
Control
del
Proceso
y
Uso
de
Plaguicidas
y
Sustancias
Tóxicas
(
CICOPLAFEST)
1998
official
catalog
as
a
restricted
pesticide,
meaning
that
a
written
recommendation
issued
by
a
technician
authorized
by
the
federal
government
is
required
for
its
non­
pharmaceutical
use.

Agricultural,
Veterinary
and
Other
Uses
Currently
lindane
is
authorized
for
use
in
Mexico
for
ectoparasite
control
on
livestock
for
ticks,
fleas,
common
fly
larvae,
etc.
It
is
also
registered
for
use
as
a
seed
treatment
for
oats,
barley,
beans,
corn,
sorghum
and
wheat,
and
as
a
soil
treatment
for
corn
and
sorghum.
Another
use
of
lindane
in
Mexico
is
listed
as
flea
treatment
for
domestic
animals.
Lindane
is
registered
in
Mexico
for
public
health
campaigns
and
was
previously
used
to
control
scorpions,
but
this
use
is
no
longer
recommended
by
the
Ministry
of
Health.

Official
information
on
amounts
of
lindane
used
for
each
purpose
is
not
available.
Based
on
information
provided
by
industry,
the
majority
of
lindane
is
used
for
agriculture
and
veterinary
uses
(
approximately
19
tons
annually).

Pharmaceutical
Uses
Pharmaceutical
uses
of
lindane
in
Mexico
include
formulation
of
creams
and
shampoos
for
scabies
and
lice
treatment.
Lindane­
containing
pharmaceutical
products
are
available
in
pharmacies
and
included
in
the
"
Cuadro
Básico
de
Salud,"
the
list
of
pharmaceuticals
required
to
be
readily
available
throughout
the
national
health
system.
6
The
estimated
amount
of
lindane
used
for
pharmaceutical
uses
is
less
than
one
ton.
Estimation
of
the
number
of
treatments
is
not
currently
available.

3.
Regulatory
Status
in
Europe
(
Draft
NARAP,
2005)

Some
countries
in
Europe
currently
allow
restricted
use
of
lindane.
In
2004,
the
European
Parliament
(
the
legislative
body
for
the
European
Union)
adopted
Regulation
(
EC,
or
European
Commission)
850/
2004
on
POPs
that
bans
the
production
and
use
of
13
intentionally
produced
POP
substances.
For
HCH/
lindane,
the
regulation
allows
member
states
a
phase
out
period
until
December
2007.
Member
states
may
request
to
use
lindane
for
professional
lumber
treatment
and
for
indoor
industrial
and
residential
applications
until
September
1,
2006.
They
may
request
to
use
lindane
for
public
health
and
as
a
chemical
intermediate
until
December
31,
2007.
For
more
information,
go
to:
http://
europa.
eu.
int/
comm/
environment/
pops/
index_
en.
htm.

Currently,
lindane
in
being
produced
in
Romania
for
the
US
technical
registrant
of
US
agricultural
products.
However,
it
is
expected
that
once
Romania
become
part
of
the
EU
in
2007,
manufacture
of
lindane
will
no
longer
be
allowed,
because
the
EU
considers
lindane
a
POP
chemical
and,
as
such,
it
is
covered
under
EC
Regulation
850/
2004
noted
above
(
Hauzenberger,
2004).

4.
International
Agreements
and
Treaties
(
Draft
NARAP,
2005)

The
Great
Lakes
Binational
Toxics
Strategy
is
a
voluntary
strategy
signed
in
1997
between
the
United
States
and
Canada
for
the
virtual
elimination
of
persistent
toxic
substances
in
the
Great
Lakes.
HCH
(
including
lindane)
is
listed
as
a
Level
II
substance,
which
means
that
either
the
US
or
Canada
has
grounds
to
indicate
its
persistence
in
the
environment,
potential
for
bioaccumulation
and
toxicity,
but
these
grounds
have
not
yet
been
sufficiently
considered
by
both
nations.
The
governments
of
Canada
and
the
United
States
encourage
pollution
prevention
activities
for
Level
II
substances
to
reduce
levels
in
the
environment
and
to
conform
to
the
laws
and
policies
of
each
country.
(
In
contrast,
Level
I
substances,
such
as
PCBs,
are
targeted
for
virtual
elimination
through
collaborative
bilateral
efforts.)
For
additional
information,
go
to:
www.
epa.
gov/
glnpo/
bns/.

The
use
of
lindane
has
been
addressed
in
at
least
two
international
treaties.
The
first
is
the
1998
Aarhus
Protocol
on
Persistent
Organic
Pollutants
(
POPs),
one
of
the
eight
protocols
under
the
Convention
on
Long­
Range
Transboundary
Air
Pollution
(
LRTAP),
negotiated
under
the
auspices
of
the
United
Nations
Economic
Commission
for
Europe
(
UNECE).
The
POPs
Protocol
entered
into
force
in
October
2003.
The
UNECE
region
includes
the
Russian
Federation,
Central
Asia,
Europe,
Canada
and
the
United
States.
HCH/
Lindane
is
one
of
the
16
POPs
substances
listed
in
this
legallybinding
Protocol.
The
Protocol
restricts
lindane
to
six
specific
uses.
As
of
September
2,
2005,
there
are
36
Signatories
and
24
Parties
to
Protocol.
Canada
is
a
Party
and
the
United
States
has
signed,
but
not
ratified
the
LRTAP
POPs
Protocol.
For
additional
7
information
on
the
LRTAP
POPs
Protocol,
go
to:
www.
unece.
org/
env/
lrtap/
pops_
h1.
htm.

In
August
2004,
Austria
prepared
a
technical
report
on
lindane,
as
part
of
a
scheduled
reassessment
under
the
Protocol
of
all
restricted
uses
of
Lindane.
For
this
report,
go
to:
www.
unece.
org/
env/
popsxg/
mtg_
tf_
pops.
htm.

The
second
international
treaty
is
the
Rotterdam
Convention
on
the
Prior
Informed
Consent
(
PIC)
Procedure
for
Certain
Hazardous
Chemicals
and
Pesticides
in
International
Trade,
which
entered
into
force
in
February
2004.
It
is
legally­
binding
for
Parties.
PIC
has
73
Signatories
and
101
Parties
as
of
January
2006.
The
PIC
includes
lindane,
reflecting
that
lindane
has
been
banned
or
severely
restricted
by
at
least
one
or
more
countries
in
two
or
more
different
regions
of
the
world.
As
of
December
2004,
26
countries
have
banned
all
import
of
lindane
and
approximately
10
have
restricted
or
severely
restricted
its
use.
Under
the
PIC,
when
an
importing
country
indicates
that
no
consent
for
import
is
provided,
exporting
countries
are
obligated
to
prevent
export
of
that
chemical
to
that
country.
The
scope
of
PIC
does
not
apply
to
pharmaceuticals,
including
human
and
veterinary
drugs.
Canada
and
Mexico
are
Parties
and
the
United
States
has
signed,
but
not
ratified
the
Rotterdam
Convention.
To
view
the
list
of
countries
that
do
not
allow
the
import
of
lindane,
go
to:
http://
www.
pic.
int.

5.
The
Stockholm
Convention
on
Persistent
Organic
Pollutants
(
POPs)
(
Draft
NARAP,
2005)

The
Stockholm
Convention
on
POPs
is
a
legally­
binding
treaty
that
calls
for
the
eventual
global
elimination
of
an
initial
list
of
12
POPs
with
specific
criteria
and
guidelines
for
adding
new
POP
substances.
There
are
specific
criteria
to
satisfy
for
adding
additional
substances
to
the
Convention,
including
persistence,
bio­
accumulation,
potential
for
long­
range
environmental
transport,
and
adverse
effects.
It
entered
into
force
in
May
2004
and
is
legally­
binding
for
Parties.
Lindane
is
not
on
the
initial
list
of
12
substances.
The
Stockholm
Convention
was
signed
by
151
nations
in
May
2001
and
has
been
ratified
by
115
nations
as
of
January
2006.
Canada
and
Mexico
are
Parties
and
the
United
States
has
signed,
but
not
ratified
the
Stockholm
Convention.
For
additional
information
on
the
Stockholm
Convention,
go
to:
www.
pops.
int.
In
June
2005,
Mexico
submitted
a
proposal
to
add
lindane
to
the
elimination
annex
to
the
Stockholm
Convention.
In
November
2005,
the
POP
Review
Committee
decided
that
the
proposal
satisfied
the
screening
criteria
and
directed
an
ad
hoc
working
group
to
prepare
a
draft
risk
profile
for
lindane
in
2006.

6.
The
North
American
Regional
Action
Plan
(
NARAP)
(
Draft
NARAP,
2005)

In
April
2000,
the
Substance
Selection
Task
Force
(
SSTF),
working
under
the
direction
of
the
Sound
Management
of
Chemicals
(
SMOC)
Working
Group
of
the
CEC,
submitted
their
conclusions
that
lindane
and
other
HCH
isomers
"
pose
risk
to
humans
and
wildlife"
in
North
America
(
see
8
http://
www.
cec.
org/
pubs_
docs/
documents/
index.
cfm?
varlan=
english&
ID=
1032).
The
SSTF
acknowledged
that
lindane
is
of
regional
concern
and
that
there
would
be
real
benefits
obtained
from
collective
action
in
the
development
and
implementation
of
a
North
American
Regional
Action
Plan
on
Lindane.
It
was
also
recommended
that
this
Action
Plan
should
identify
issues
related
to
key
implementation
measures.

Following
these
recommendations,
in
July
2002,
the
CEC
Council
of
Ministers
issued
Resolution
02­
07
directing
the
SMOC
Working
Group
to
develop
a
NARAP
on
lindane.
Further
information
is
available
at:
www.
cec.
org/
programs_
projects/
pollutants_
health/
smoc/.

The
initial
draft
of
the
Lindane
NARAP
was
developed
in
September
2004.
The
draft
has
gone
through
intra­
and
inter­
agency
review
in
all
three
countries
and
through
a
public
comment
period
that
ended
in
November
2005.
The
final
intra­
and
inter­
agency
review
began
on
January
3,
2006
and
will
end
in
mid­
February
2006.
The
goal
is
to
have
the
NARAP
finalized
by
early
April
2006
so
that
it
can
be
transmitted
to
the
CEC
Council
for
approval
in
early
June
2006.
It
is
expected
that
Mexico
will
assume
the
chairmanship
of
the
Lindane
Task
Force
for
the
implementation
phase
of
the
NARAP.
9
II.
Properties
and
Production
of
Hexachlorocyclohexanes
(
HCH)

A.
Chemical
Identity
Lindane
(
gamma
HCH)
and
other
HCH
isomers
are
in
a
family
of
manufactured
substances
known
as
organochlorine
chemicals.
In
this
case,
the
HCH
isomers
have
one
basic
chemical
structure,
1,2,3,4,5,6­
C6H6Cl6,
but
the
chlorine
atoms
are
found
in
varying
orientations
in
the
molecule.
Depending
on
the
orientation
of
the
chlorine
atoms,
whether
being
axial
(
a)
or
equatorial
(
e),
these
isomers
are
named
 ­
(
exists
in
(+)
and
(­)
enantiomeric
forms),
 ­,
 ­,
 ­
and
  
HCH.
The
physical
properties
and
the
environmental
fate
behaviors
of
each
isomer
differ
because
of
the
different
chlorine
atom
orientations
on
each
molecule
(
see
Figure
1
below).
Selected
physical
and
chemical
properties
of
 ­,
 ­,
and
 ­
HCH
isomers
are
listed
in
Table
1.

HCH
was
first
synthesized
in
1825
by
reaction
of
benzene
with
chlorine
in
the
presence
of
sunlight
(
ultraviolet­
radiation)
to
produce
what
was
then
called
BHC
or
"
benzene
hexachloride."
This
terminology
is
no
longer
used.
Current
nomenclature
refers
to
technical
HCH,
a
mixture
of
all
6
stable
isomers
that
was
used
as
a
pesticide
prior
to
the
isolation
of
the
only
HCH
isomer
that
exhibits
insecticidal
properties,
gamma
HCH
or
lindane.
The
insecticidal
properties
of
technical
HCH
were
first
described
in
the
1940s.
In
insects,
lindane
acts
through
the
inhibition
of
the
gamm
  
aminobutyric
acid
(
GABA)
receptor
of
the
central
nervous
system
(
CNS).
GABA
operates
by
increasing
chloride
ion
permeability
into
neurons,
thereby
inhibiting
neurostimulation
inducing
overstimulation
of
the
CNS
causing
rapid
violent
convulsions.
The
alpha
isomer
is
much
less
active
at
inhibiting
binding
to
the
GABA
receptor
than
lindane
and
the
beta
isomer
seems
not
to
exhibit
inhibiting
binding
at
all.

HCH
has
been
commercialized
in
two
predominant
products:
technical
HCH
and
purified
gamma
isomer,
lindane.
At
one
point,
it
was
also
sold
as
fortified
BHC
with
47%
gamma
isomer.
Technical
HCH
contains
about
60­
70%
alpha­
HCH,
5­
12%
beta­
HCH,
and
10­
15%
gamma­
HCH,
or
lindane.
These
are
the
three
most
environmentally
significant
isomers.
10
Copyright
©
2005
American
Chemical
Society
Figure
1.
Structures
of
the
 ­,
 ­,
and
 ­
isomers
of
hexachlorocyclohexane.
 ­
HCH
has
one
plane
of
symmetry,
and
 ­
HCH
has
three
planes
of
symmetry,
whereas
 ­
HCH
is
a
chiral
molecule.
The
diagrams
on
the
right
indicate
the
chlorine
substitutions
in
the
axial
(
a)
and
equatorial
(
e)
orientations
(
Reprinted
with
permission
from
Xiao
et
al.,
2004).

Table
1.
Selected
physical
and
chemical
properties
of
 ­,
 ­,
and
 ­
HCH
isomers
Property
(
units)
 ­
HCH
 ­
HCH
 ­
HCH
CAS
registry
319­
84­
6
319­
84­
7
58­
89­
9
Chemical
Formula
C6H6Cl6
C6H6Cl6
C6H6Cl6
Molecular
Weight
290.85
290.85
290.85
Physical
State
Crystalline
solid
Crystalline
solid
Crystalline
solid
Solubility
in
Water
(
ppm)
10.0
5.0
7.3
Log
Kow
3.8
3.78
3.72
Log
Koc
3.57
3.57
3.0­
3.57
11
Table
1.
Selected
physical
and
chemical
properties
of
 ­,
 ­,
and
 ­
HCH
isomers
Property
(
units)
 ­
HCH
 ­
HCH
 ­
HCH
Vapor
Pressure
(
mm
Hg)
4.5
x
10­
6
@
25
B
C
3.6
x
10­
7
@
20
B
C
4.2
x
10­
6
@
20
B
C
Henry's
law
Constant
(
atm­
m3/
mol)
6.68
x
10­
6
@
25
B
C
6.9
x
10­
7
@
20
B
C
3.5
x
10­
6
@
25
B
C
B.
Manufacturing
Process
Lindane
and
its
precursor
technical
hexachlorocyclohexane,
or
technical
HCH,
do
not
occur
as
natural
substances.
Technical
HCH
is
manufacturied
by
a
chemical
reaction
(
the
photochlorination
of
benzene)
that
produces
a
mixture
of
the
different
isomers
of
hexachlorocyclohexane
(
Johnson,
2005).
These
isomers
and
their
typical
yield
are
listed
in
Table
2
below:

Table
2:
Ratio
of
Isomers
in
the
Production
of
Technical
HCH
(
Draft
NARAP,
2005)
HCH
Isomer
Percent
in
synthesis
mixture
Alpha­
HCH
60
­
70
Beta­
HCH
5­
12
Gamma­
HCH
(
Lindane)
10­
15
Delta­
HCH
6­
10
Epsilon­
HCH
3­
4
Lindane
(
gamma­
HCH)
is
then
extracted
from
this
mixture
and
purified.
99%
pure
lindane
is
produced
at
a
10­
15
percent
yield
from
technical
HCH;
therefore,
for
every
ton
of
lindane
that
is
produced,
approximately
6
­
10
tons
of
other
isomers
are
produced.
The
other
HCH
isomers
do
not
have
the
insecticidal
properties
of
the
gamma
isomer,
and
no
other
use
has
been
found
for
these
isomers.

Diagram
1
below
further
illustrates
the
lindane
manufacturing
process
and
the
ultimate
potential
environmental
release
of
lindane
and
the
other
isomers.
12
Diagram
1.
Lindane
­
Manufacture,
Formulation,
Use,
Disposal,
and
Release
Benzene
+
Cl2
mixed
isomers
of:
Hexachlorocyclohexane
(
HCH)
uv
solvent
distillation
fractional
crystallization
gamma
HCH
(
Lindane)
[
12%]

Formulate
w/
inerts
&
six
other
pesticides
Lindane
Pesticide
Products
seed
treatment
Treated
Seeds
seed
planting
Farm
Field
volatilization
run
off
degradation
Other
Isomers
[>
85%]
alpha
HCH
(
65
%)
beta
HCH
(
10
%)
delta
HCH
(
8
%)
epsilon
HCH
(
3
%)
others
(
1
%)
>
85
%

cracking
disposal
volatilization
Landfill
run
off
degradation
Trichlorobenzenes
HCl
Others
(
possible
dioxins)

Possible
Market
Possible
incineration
(
potential
dioxins)

Possible
incineration
(
potential
dioxins)
13
For
decades
these
waste
isomers
were
generally
disposed
of
in
open
landfills,
often
in
fields
and
other
disposal
sites
near
the
HCH
manufacturing
facilities.
After
disposal,
degradation,
volatilization,
and
run
off
of
the
waste
isomers
occur.
The
technical
registrant
has
claimed
in
its
comments
on
the
draft
NARAP
(
Johnson,
2005)
that
modern
production
processes
further
treat
the
waste
isomers
(
this
treatment
is
known
as
"
cracking")
to
produce
other
chemicals,
trichlorobenzene
(
TCB)
and
hydrochloric
acid
(
HCl),
that
can
be
sold.
While
the
Agency
does
not
doubt
that
cracking
of
the
waste
isomers
is
possible,
the
Agency
cannot
confirm
that
the
cracking
process
is
actually
occurring
at
lindane
manufacturing
facilities
(
Jensen,
2004).
The
Agency
also
cannot
ensure
that
the
waste
isomers
are
treated
since
the
manufacturing
of
lindane
occurs
outside
of
the
US.
In
addition,
there
is
the
potential
for
dioxin
formation
during
the
cracking
process
or
during
the
incineration
of
waste
isomers
or
the
products
of
cracking
(
Jensen,
2004).
Therefore,
the
Agency
is
not
certain
that
cracking
is
a
viable
and
practiced
method
for
handling
waste
isomer
formation
during
the
manufacture
of
lindane.

C.
Global
Use
and
Production
of
HCH
Global
Use
of
HCH
(
Li
and
Vijgen,
2005).

Historical
useage
of
lindane
worldwide
for
agricultural
purposes
between
1950
and
2000
was
approximately
450
kilo
tons
(
kt),
of
which
280
kt
was
used
in
Europe,
73
kt
was
used
in
Asia,
and
64
kt
was
used
in
America.
Global
annual
lindane
use
was
highest
in
the
1960s
and
early
1970s,
and
has
since
been
declining;
global
use
in
2000
was
approximately
2
kt,
or
2204
tons.
Lindane
has
also
been
used
on
livestock,
forestry,
human
health,
and
for
other
purposes;
therefore,
the
total
global
lindane
used
for
all
purposes
could
be
as
high
as
approximately
600
kt.

Total
lindane
usage
in
the
US
between
1950
and
2004
was
18.8
kt.
Current
estimates
of
annual
US
use
of
lindane
range
from
150
tons
(
Li
and
Vijgen,
2005)
to
75
tons
(
Brassard,
2005).

Since
2000,
a
number
of
regulatory
actions
have
been
initiated
that
have
affected
and
are
expected
to
continue
to
affect
the
global
demand
for
lindane
(
see
Section
I.
B
of
this
document).
For
instance,
since
2004,
Europe
has
banned
the
production
of
lindane,
and
use
of
lindane
is
being
phased
out
by
December
2007;
Canada
has
been
phasing
out
lindane
agricultural
use
since
1999;
and,
as
of
2005,
Mexico
is
also
eliminating
all
lindane
use.
As
such,
the
percentage
of
global
lindane
production
occurring
due
to
US
seed
treatment
registrations
of
lindane
has
increased
in
recent
years.

According
to
Li
and
Vijgen
(
2005),
estimating
global
usage
of
lindane
over
time
is
difficult
because
many
countries
do
not
keep
records
of
or
have
no
reporting
mechanism
for
information
on
pesticide
production,
sales,
and/
or
usage,
and
other
countries
consider
this
information
proprietary.
14
North
American
Production
of
HCH
Lindane
is
no
longer
manufactured
in
North
America.
Lindane
was
manufactured
in
the
United
States
until
1977,
however,
official
records
are
sparse
to
non­
existent.
Information
on
a
former
lindane
manufacturing
site
in
Nevada
illustrates
the
scale
of
the
waste
isomer
problem.
According
to
the
Nevada
Department
of
Environmental
Protection,
a
company
manufactured
approximately
12,000
tons
of
lindane,
and
approximately
50,000
tons
of
waste
HCH
isomers
have
been
buried
at
the
site
since
the
late
1970s
and
capped
with
a
clay
liner.
1
Lindane
and
HCH
are
subject
to
the
disposal
requirements
of
the
Resource
Conservation
and
Recover
Act
(
RCRA).
The
disposal
techniques
over
the
years
prior
to
the
RCRA
requirements
have
lead
to
dozens
of
sites
with
potential
uncontrolled
HCH
residue
releases.
Many
of
these
older
disposal
sites
are
now
subject
to
requirements
of
Comprehensive
Environmental
Response,
Compensation,
and
Liability
Act
(
CERLA,
commonly
referred
to
as
"
Superfund").
At
least
189
sites
in
the
US
have
one
or
more
of
the
HCH
isomers
and
have
been
identified
for
the
National
Priorities
List
(
NPL).
The
NPL
is
the
list
of
national
priorities
among
the
known
releases
or
threatened
releases
of
hazardous
substances,
pollutants,
or
contaminants
throughout
the
US
and
its
territories.
(
The
NPL
is
intended
primarily
to
guide
the
EPA
in
determining
which
sites
warrant
further
investigation.)

Global
Production
of
HCH
Currently,
for
the
world
market,
lindane
is
produced
in
India,
Romania
and
possibly
Russia.
China
is
reported
to
have
been
the
major
world
producer
of
technical
HCH,
accounting
for
more
than
4.5
million
metric
tonnes
(
about
5
million
tons)
between
1945
and
1983
(
Draft
NARAP,
2005).
In
1983,
China
banned
both
the
use
of
technical
HCH
and
production
of
technical
HCH,
except
as
an
intermediary
in
the
production
of
lindane.
Manufacture
of
lindane
in
China
stopped
in
2003
(
Li,
2005;
Peking
University,
2005)

According
to
the
US
technical
registrant,
lindane
is
produced
in
Romania
for
agricultural
products
used
in
the
US.
No
information
is
available
on
the
amounts
of
lindane
produced
or
used
in
Romania.
In
addition,
there
is
no
historical
information
on
the
amounts
of
HCH
and/
or
lindane
produced
in
India;
however,
the
Agency
understands,
based
on
personal
communications,
that
there
is
at
least
one
company
that
currently
produces
lindane
(
Jensen,
2004).
It
was
communicated
that,
because
of
the
drop
in
demand,
this
company
is
producing
only
300
kg
of
lindane
per
day,
six
months
per
year,
and
the
company
reported
no
production
in
2004.
In
2003,
the
plant
built
a
land
fill
to
cap
the
estimated
3,000
tons
of
waste
isomers
(
Jensen,
2004).
The
Agency
has
no
information
on
historical
technical
HCH
production
information
from
the
former
Soviet
Union
or
Russia.

1
State
of
Nevada,
2004
Personal
Communication
between
Todd
Croft,
State
of
Nevada
Division
of
Environmental
Protection,
Las
Vegas
office,
and
Janice
Jensen,
USEPA
Office
of
Pesticide
Programs,
November
17,
2004.
15
III.
HCH
Environmental
Fate
and
Effects
Risk
Assessment
The
Agency
conducted
an
environmental
fate
and
effects
risk
assessment
to
support
the
2002
RED
for
lindane.
The
lindane
risk
assessment
for
the
RED
can
be
found
at
www.
regulations.
gov
.

The
Agency
initially
did
not
assess
exposures
to
other
HCH
isomers
as
part
of
its
Environmental
Fate
and
Effects
Assessment.
However,
environmental
and
other
organizations
representing
indigenous
people
submitted
comments
advocating
that
EPA
consider
exposure
to
other
HCH­
isomers,
particularly
$­
HCH,
as
a
result
of
lindane
use
(
Docket
Control
Number
OPP­
2002­
0202).

A.
Environmental
Fate
and
Transport
1.
Persistence
The
ubiquity
of
HCH
isomers
in
the
various
environmental
compartments
has
resulted
from
the
extensive
use
as
an
insecticide,
as
well
as
during
manufacturing
and
disposal
of
these
chemicals.
Once
released
into
the
environment,
the
primary
process
of
dissipation
of
HCHs
from
soil
is
volatilization
into
the
air,
although
abiotic
and
biotic
degradation
as
well
as
uptake
by
crops
can
also
occur.
However,
these
isomers
are
resistant
to
abiotic
processes
like
photolysis
and
hydrolysis
(
except
at
high
pH),
and
degrade
very
slowly
by
microbial
actions.
The
structure
of
 ­
HCH
confers
a
greater
stability
as
compared
with
 ­
and
 ­
HCH,
and
reflected
in
its
environmental
and
biological
persistence
(
Willett
et
al.,
1998
and
Xiao
et
al.,
2004).
Since
most
degradation
pathways
occur
slowly,
the
presence
of
the
degradates
is
generally
at
relatively
low
levels.
Possible
degradates
could
include
pentachlorocyclohexene,
1,2,4,­
trichlorobenzene,
and
1,2,3­
trichlorobenzene
(
U.
S.
EPA,
2001).

a.
Abiotic
Degradation
and
Transformation
Hydrolysis
Hydrolysis
is
not
considered
an
important
degradation
process
for
HCH
isomers
in
aquatic
environments
under
neutral
pH
conditions.
The
hydrolysis
half­
lives
of
lindane
in
solution
were
reported
to
be
$
732
days
to
stable
at
pH
5
and
pH
7,
and
43
to
182
days
at
pH
9
(
U.
S.
EPA,
2001
and
Das,
1990).
Hydrolysis
rate
constants
for
 ­
and
 ­
HCH
have
been
determined
as
a
function
of
temperature
and
pH
(
Ngabe
et
al.,
1993).
The
estimated
hydrolysis
half
lives
for
 ­
and
 ­
HCHs
in
seawater
were
0.8
and
1.1
years
at
pH
8
and
20
B
C,
respectively.
The
hydrolysis
rates
decline
with
temperature,
and
estimated
half­
lives
for
Lake
Huron
at
pH
7.8
and
5
B
C
were
26
years
for
 ­
HCH
and
42
years
for
for
 ­
HCH.
Based
on
these
rate
constants,
the
hydrolysis
rates
were
estimated
to
be
even
slower
in
the
Arctic
Ocean
at
pH
8
and
0
B
C.
The
calculated
half
lives
were
63
years
for
 ­
HCH
and
110
years
for
 ­
HCH
(
Harner
et
al.,
1999).
16
Photolysis
in
Air
Since
HCH
does
not
absorb
light
>
290
nm,
direct
photolysis
in
the
troposphere
is
not
expected
to
be
an
important
environmental
fate
process.
However,
HCH
may
degrade
in
the
atmosphere
by
reacting
with
photochemically
produced
hydroxyl
radicals.
The
rate
constants
for
the
reaction
of
 ­
HCH
and
 ­
HCH
with
hydroxyl
radicals
were
measured
as
1.9x10­
13
and
1.4x10­
13
cm/
molecule­
second,
respectively
(
Brubaker
and
Hites
1998).
Calculation
based
on
these
rate
constants
and
tropospheric
average
hydroxyl
radical
concentration
gave
estimated
lifetimes
in
air
of
120
days
for
 ­
HCH
and
96
days
for
 ­
HCH.
In
locations
where
the
atmospheric
hydroxyl
radical
concentration
is
very
low,
the
persistence
times
of
these
compounds
are
much
longer.
Cortes
and
Hites
(
2000)
estimated
that
the
average
half­
life
of
 ­
HCH
and
 ­
HCH
around
the
Great
Lakes
region
ranged
from
about
3
to
4
years.
However,
Chen
et
al.
(
1984),
reported
photodegradation
half­
lives
of
91,
152,
and
104
hours
for
thin
films
of
 ­
HCH,
 ­
HCH,
and
 ­
HCH,
respectively
when
irradiated
with
light
of
wavelength
295 
305
nm.
No
absorption
bands
were
observed
in
this
spectral
region.
Therefore,
the
mechanism
of
photodegradation
of
these
HCH
isomers
and
its
environmental
significance
are
uncertain.

Photolysis
in
Water
Somewhat
conflicting
information
is
available
on
the
rate
of
photolysis
of
 ­
HCH
in
water.
Since
HCH
does
not
contain
chromophores
that
absorb
light
>
290
nm,
direct
photolysis
is
not
expected
to
occur.
However
indirect
photolysis,
whereby
a
photosensitizing
agent
may
absorb
light
and
then
transfer
its
excitation
energy
to
HCH,
may
occur.
Humic
and
fulvic
acids
are
well­
known
photosensitizing
agents
and
are
practically
ubiquitous
in
natural
waters.
Saleh
et
al.
(
1982)
reported
that
 ­
HCH
firstorder
photolysis
half­
lives
of
169,
1,791,
and
1,540
hours
in
pond
water,
lake
water,
and
water
from
a
quarry
at
pH
9.3,
7.3,
and
7.8,
respectively
when
solutions
were
exposed
to
direct
sunlight.
However,
the
rapid
rate
of
degradation
at
pH
9.3
may
have
been
enhanced
by
hydrolysis
reactions
rather
than
by
photolysis.
Oxidants
commonly
found
in
natural
waters,
such
as
peroxy
radicals,
hydroxyl
radicals,
and
singlet
oxygen
species,
can
degrade
HCH
in
water.
Mill
(
1999)
estimated
that
the
indirect
photolysis
half­
life
of
HCH
in
natural
waters
is
about
270
days,
and
the
dominant
oxidant
for
HCH
was
the
hydroxyl
radical.
Photolysis
of
 ­
HCH
in
aqueous
solution
in
the
presence
of
polyoxomethallate,
a
strong
oxidizing
agent,
has
also
been
demonstrated
(
Hiskia
et
al.
1997).
However,
USEPA
(
2001)
reported
that
there
is
no
evidence
of
aqueous
photodegradation
of
lindane
during
a
30­
day
study
period,
even
when
acetone
was
used
as
a
photosensitizer.

Photolysis
in
Soil
As
mentioned
earlier,
direct
photolysis
in
not
expected
as
a
route
of
transformation
for
HCH
isomers,
although
indirect
photolysis
may
occur
in
presence
of
photosensitizer
like
humic
substance
in
soil.
On
a
1­
mm
thick
soil
specimen
exposed
to
artificial
sunlight
for
12
hour
per
day,
lindane
degraded
only
very
slightly
over
the
30­
day
test
period.
The
extrapolated
half­
life
was
greater
than
150
days
(
U.
S.
EPA,
2001).
17
The
dark
control
showed
a
5%
loss
over
the
30­
day
study.
The
soil
degradation
half­
life
with
consideration
for
the
dark
control
losses
is
200
days.
Because
of
the
extreme
extrapolation
to
obtain
a
half
life,
this
study
essentially
provides
no
evidence
of
lindane
photodegradation
on
soil.

b.
Biotic
Degradation
and
Transformation
In
Soils
and
Sediments
In
a
336­
day
aerobic
soil
metabolism
study
under
a
laboratory
conditions,
lindane
degraded
very
slowly,
with
a
calculated
half
life
of
980
days
(
U.
S.
EPA,
2001).
The
major
degradation
product
was
pentachlorocyclohexane
(
PCCH),
which
reached
maximums
of
3.84%
of
applied
lindane.
In
a
silt
loam
soil/
corncob
test
matrix,
34.7%
of
the
compound
was
degraded
over
a
60­
day
test
period,
whereas
53.5%
degradation
was
observed
in
liquid
cultures
over
a
30­
day
test
period
(
Kennedy
et
al.
1990).
The
results
of
this
study
have
been
confirmed
by
more
recent
studies
(
Mougin
et
al.,
1996).
The
isolation
of
 ­
HCH­
degrading
bacteria,
classified
as
Sphingomonas
paucimobilis,
from
contaminated
soils
has
been
reported
(
Thomas
et
al.
1996).
A
Pseudomonas
species
has
also
been
isolated
from
pretreated
soil
that
is
able
to
degrade
 ­
HCH
and
 ­
HCH,
but
not
 ­
HCH,
within
10 
20
days
under
both
saturated
(
anaerobic)
and
unsaturated
(
aerobic)
conditions;
greater
degradation
rates
were
observed
under
aerobic
conditions
(
Sahu
et
al.
1993).
The
presence
of
crops
on
the
soils
also
affects
the
persistence
of
HCH
residues,
with
half­
lives
of
58.8
and
83.8
days
for
cropped
and
uncropped
plots,
respectively.
 ­
HCH
was
the
most
persistent
isomer,
with
half­
lives
of
184
and
100
days,
respectively,
on
cropped
and
uncropped
plots;
 ­
HCH
was
next
at
107
and
62.1
days,
followed
by
 ­
HCH
at
54.4
and
56.1
days.

In
Water
The
primary
dissipation
process
of
HCHs
from
soil
and
water
is
volatilization
into
the
atmosphere,
although
biotic
degradation
can
also
occur
to
a
lesser
extent.
However,
the
rates
of
degradation
depend
on
the
ambient
environmental
conditions.
Microbial
degradation
proceeds
faster
than
hydrolysis,
especially
in
cold
environments.
Harner
et
al.
(
1999
and
2000)
estimated
half­
lives
in
eastern
Arctic
Ocean
water
of
6
years
for
(+)
enantiomer
and
23
years
for
(­)
enantiomer
of
 ­
HCH
and
19
years
for
 ­
HCH.
The
halflives
of
(+)
and
(­)
enantiomers
of
 ­
HCH
in
a
small
Arctic
lake
were
estimated
to
be
0.6
and
1.4
years
respectively
(
Helm
et
al.,
2000).
Zoetemann
et
al.
(
1980)
estimated
the
degradation
half­
lives
of
3
days
for
river,
30
days
for
lake,
and
>
300
days
for
groundwater
for
lindane.
These
half­
lives
are
much
shorter
than
hydrolytic
half­
lives
described
earlier.
Harner
et
al.
(
2000)
estimated
that
microbial
degradation
could
account
for
about
37%
of
 ­
HCH
and
29%
of
 ­
HCH
of
the
annual
loss
from
the
Arctic
Ocean.
There
is
no
information
on
 ­
HCH.
18
2.
Isomerization
of
the
HCHs
Oehme
(
1991)
and
Patton
et
al.
(
1989)
have
attributed
the
isomerization
of
lindane
in
contributing
high
concentrations
of
 ­
HCH
in
the
Arctic.
Several
laboratory
experiments
have
demonstrated
the
conversion
of
lindane
to
other
isomers
via
photoisomerization
in
air
(
Steinwander,
1976;
Hamada
et
al.
1981;
Malaiyandi
and
Shah
1984)
and
bio­
isomerization
in
water,
soil
and
sediments
(
Benezet
et
al.,
1973;
Huhnerfuss
et
al.,
1992
).
Walker
et
al.
(
1999)
reviewed
many
laboratory
and
field
studies
to
explore
the
potential
of
lindane
conversion
into
other
isomers
and
to
corroborate
with
the
spatial
and
temporal
monitoring
data
of
 ­
and
 ­
HCHs.
They
concluded
that
even
though
the
laboratory
data
suggest
lindane
may
transform
into
other
isomers,
air
monitoring
data
do
not
support
the
evidence
of
significant
isomerization
of
HCHs.
Walker
et
al.
(
1999)
noted
that
if
photochemical
transformation
of
 ­
HCH
to
 ­
HCH
in
air
takes
place,
one
should
see
significant
concentrations
of
 ­
HCH
in
the
Southern
Hemisphere
air.
A
considerable
reduction
in
 ­
HCH
concentrations
has
been
observed
recently
in
Arctic
air
(
Shen
et
al.,
2005
and
Li
et
al.,
1998).
Furthermore,
Walker
et
al.
(
1999)
note
that
reported
concentrations
of
 ­
HCH
in
the
air
of
southern
Norway
dropped
by
50%
between
1991
and
1995
(
Haugen
et
al.,
1998),
but
no
such
pattern
was
observed
for
 ­
HCH.
Recent
studies
also
indicated
that
bio­
isomerization
may
play
an
insignificant
role
in
the
overall
degradation
of
 ­
HCH
(
Buser
and
Muller,
1995;
Walliszewski
1993;
Singh
et
al.,
1991).
Walker
et
al.
(
1999)
concluded
that
there
is
no
significant
evidence
of
isomerization
of
lindane
to
contribute
higher
 ­
HCH
concentrations
in
the
environment.

3.
Mobility
in
the
Environment
Once
released
into
the
environment,
HCH
isomers
can
partition
into
various
environmental
media.
The
HCH
isomers
present
in
soil
can
leach
to
groundwater,
sorb
to
soil
particulates
and
transport
to
surface
water
via
runoff,
or
volatilize
to
the
atmosphere.
However,
Henry's
Law
Constant
(
Table
2
in
Section
II
of
this
document)
of
these
isomers
suggest
that
volatilization
is
the
most
important
route
of
dissipation
from
water
and
moist
soils
followed
by
aerial
long­
range
transport.
Adsorption
of
HCH
isomers
to
soil
and
sediments
is
generally
a
preferential
partitioning
process
after
volatilization.
Leaching
of
HCH
isomers
through
soil
is
governed
by
their
water
solubility
and
their
propensity
to
bind
to
soil,
and
these
isomers
are
reasonably
soluble
in
water.
The
calculated
Koc
of
lindane
ranged
from
942
to
1798
mL/
g,
with
a
mean
of
1368
mL/
g
for
four
soils
tested
(
U.
S.
EPA,
2001).
These
data
suggest
that
lindane
has
low
leaching
potential.
Data
also
indicate
that
lindane
is
expected
to
adsorb
to
suspended
solids
and
sediment
in
water.
Based
on
the
results
of
a
number
of
laboratory
soil
column
leaching
studies
that
used
soils
of
both
high
and
low
organic
carbon
content
as
well
as
municipal
refuse,
 ­
HCH
generally
has
low
mobility
in
soils
(
Melancon
et
al.
1986;
Reinhart
et
al.
1991).
In
a
study
involving
a
laboratory
sediment/
water
system
(
pH=
7.42;
2.18%
organic
carbon)
 ­,
 ­,
and
 ­
HCH
isomers
were
highly
adsorbed
on
sediments
under
both
aerobic
and
anaerobic
conditions
and
few
differences
were
noted
in
the
adsorption
behavior
of
each
isomer
(
Wu
et
al.
1997).
The
 ­
HCH
isomer
comprised
80 
100%
of
the
total
HCH
residues
found
in
soil
or
vegetation
on
land
surrounding
an
industrial
landfill
in
Germany
10
years
after
the
final
HCH
input
(
Heinisch
et
al.
1993).
19
4.
Bioaccumulation,
Biomagnification,
and
Bioconcentration
The
octanol­
water
partition
coefficient
(
log
Koc
$
3.70,
see
also
Table
2)
for
HCH
isomers
indicates
that
they
have
the
potential
to
bioaccumulate.
The
behavior
of
HCH
isomers
in
the
environment
is
complex
because
they
are
multimedia
chemicals,
existing
and
exchanging
among
different
compartments
of
the
environment
such
as
atmosphere,
surface
water,
soil
and
sediment.
In
addition,
temperature,
humidity,
and
other
environmental
properties
may
have
significant
influence
on
environmental
degradation
rates.
The
physical
and
chemical
properties
of
the
HCH
isomers
(
Table
2)
can
be
quite
different
from
one
another.
For
example,
 ­
HCH
has
a
lower
vapor
pressure
and
a
higher
bio­
concentration
factor
in
fat
than
either
 ­
HCH
or
 ­
HCH.
In
contrast,
lindane
and
 ­
HCH
seem
to
be
more
volatile
than
 ­
HCH.
These
properties
likely
reflect
some
of
the
differences
seen
in
HCH
isomer
persistence
and
variability
in
biomagnification,
bioconcentration
and
bioaccumulation
in
the
various
biological
compartments.
Differences
in
accumulation
are
also
likely
due
to
different
modes
of
uptake,
metabolism
and
sources
of
contamination.

Bio­
concentration
factors
(
BCF)
for
lindane
were
780x
in
fillet,
2500x
in
viscera
and
1400x
in
whole
fish.
It
would
seem
this
is
partly
due
to
high
lipid
solubility.
Lindane
can
become
enriched
in
lipid­
containing
biological
compartments.
However,
although
lindane
may
bioconcentrate
rapidly,
most
data
suggest
bio­
transformation,
depuration
and
elimination
are
relatively
rapid
once
exposure
is
eliminated.
After
a
28
day
exposure
and
14
days
of
depuration,
levels
were
reduced
by
96%,
95%
and
85%
in
fillet,
viscera
and
in
whole
fish,
respectively
(
U.
S.
EPA,
2001).
The
BCF
of
 ­
HCH
(
1520
±
276)
was
significantly
higher
than
those
of
the
 ­
HCH
(
1100
±
175)
and
 ­
HCH
(
920
±
131)
isomers,
using
zebra
fish
under
a
steady­
state
condition
(
Butte
et
al.,
1991).
Geyer
et
al.
(
1997)
reported
that
the
mean
bioconcentration
factor
of
lindane
on
a
lipid
basis
(
BCFL)
is
11000
in
aquatic
organisms.

HCH
bio­
accumulation/
food
chain
data
from
Russia
(
Moisey
et
al.,
2001)
and
from
Central/
Western
Canada
(
Kelly
and
Gobas
2001)
suggests
that
 ­
HCH
bioaccumulates
bio­
magnifies,
as
does
  
HCH
(
but
at
a
lower
level),
and
lindane
biomagnifies
bio­
accumulates
the
least.
Data
from
Moisey
et
al.
(
2001)
suggest
that
the
relative
proportions
of
HCH
isomers
varied
dramatically
across
species
in
the
Arctic
marine
food
web
studied.
Kelly
and
Gobas
(
2001)
indicate
that
the
fugacity
of
lindane
decreases
with
increasing
trophic
level
suggesting
trophic
dilution
(
the
lichen­
caribouwolf
food
chain
was
studied).
It
appears
that
upper
trophic
level
mammals
may
be
able
to
efficiently
eliminate
lindane
and
to
a
smaller
extent
 ­
HCH,
but
not
 ­
HCH.
In
birds,
 ­
HCH
seems
to
have
a
tendency
to
accumulate
to
a
greater
extent
than
the
other
isomers,
which
may
be
due
to
consuming
contaminated
prey
(
Elliot
et
al.,
1989),
although
concentrations
have
been
on
the
decline.
Even
though
concentrations
of
HCH
isomers
were
detected
in
surface
waters
of
the
Arctic,
bioaccumulation
in
the
aquatic
food
chains
was
significantly
less
than
the
other
organochlorine
compounds
analyzed
(
Norstrom
and
Muir,
1994)
analyzed.
20
Although
there
is
evidence
that
HCH
isomers
can
and
do
biomagnify,
bioconcentrate
and
bioaccumulate
in
different
biological
compartments
and
at
different
rates,
the
overall
magnitude
is
an
uncertainty.
Overall,
lindane
seems
to
accumulate
environmentally,
but
generally
to
a
lesser
extent
than
either
 ­
HCH
or,
especially
 ­
HCH.
Generally,
lindane
tends
to
bio­
magnify
in
lower
trophic
levels
where
biotransformation
was
minimal,
although
not
to
the
extent
 ­
HCH
does.
 ­
HCH
tends
to
mainly
bio­
accumulate
in
upper
trophic
levels
(
fish,
birds,
mammals)
at
higher
concentrations.

5.
Presence
of
HCH
Isomers
in
the
Environment
Due
to
extensive
use
over
the
past
50
years,
the
recalcitrant
nature,
and
longrange
transport,
HCH
isomers
have
been
detected
in
air,
surface
water,
groundwater,
sediment,
soil,
ice,
snowpack,
fish,
wildlife
and
humans.
The
most
common
isomers
found
in
the
environment
are
 ­,
 ­,
and
 ­
HCHs,
with
 ­
HCH
as
the
predominant
isomer
in
air
and
ocean
water,
and
 ­
HCH
the
predominant
isomer
in
soils,
animal
tissues
and
fluids
(
Macdonald
et
al,
2000
and
Willett
et
al.,
1998).
Generally,
the
concentrations
of
HCH
isomers
in
air
and
water
are
higher
in
the
northern
hemisphere
than
the
southern
hemisphere
(
Iwata.
et
al.,
1993).
Increasing
concentration
gradient
with
latitude
reflect
the
global
condensation
of
HCH
isomers
in
the
colder
regions.
A
latitudinal
gradient
of
 ­
and
 ­
HCH
was
observed
in
a
study
examining
residue
in
tree
bark
(
Simonich
and
Hites,
1995).
A
similar
gradient
of
  
HCH
was
also
observed
in
seawater
(
Wania
and
Mackey,
1995).
In
a
global
survey,
the
detection
of
HCH
isomers
in
the
surface
waters
were
highest
(
average
of
1000
pg/
L)
in
latitudinal
waters
north
of
40
B
N
such
as
Chukchi
Sea,
Bering
Sea,
Gulf
of
Alaska,
and
north
Pacific
(
Iwata.
et
al.,
1993).
Chernyak
et
al.,
(
1995)
also
observed
a
similar
trend
of
increasing
 ­
HCH
concentrations
in
surface
water
that
ranges
from
810
to
1200
pg/
L
on
a
transect
from
the
Sea
of
Japan
to
the
Bering
Sea,
which
also
suggest
increasing
the
concentrations
of
HCH
isomers
with
increasing
latitude.
Total
HCH
residues
were
3100
pg/
L
for
Bering
Sea
and
3600
pg/
L
for
the
Chukchi
Sea,
and
the
 ­
HCH
made
up
13
to
15%
of
the
total
concentration
in
these
water
samples.

HCH
isomers
(
mainly
 
and
 )
were
the
major
organochlorine
insecticide
detected
in
Arctic
air,
snow
and
seawater
(
Barrie
et
al.
1992).
The
Arctic
is
considered
a
"
sink"
for
persistent
organic
pollutants.
Pathways
of
HCHs
to
the
Arctic
involve
not
only
transport
through
the
atmosphere
but
also
via
ocean
currents.
Ocean
transport
seems
especially
important
in
the
past
for
 ­
HCH,
which
has
the
highest
gas­
phase
solubility
in
water
(
Li
et
al.,
2002).
Li
and
Macdonald
(
2005)
estimated
the
loading
of
 ­
HCH
using
a
mass
balance
model
for
the
Arctic
Ocean
for
a
period
of
1991
to
2000
and
concluded
that
ocean
current,
atmospheric
deposition,
and
rivers
supplied
50%,
40%
and
10%
of
loadings
respectively.
They
also
concluded
that
the
temporal
and
spatial
variations
in
the
Arctic
air
and
waters
also
reflect
the
HCHs
residues
in
marine
organisms.

Although
 ­
HCH
concentrations
are
higher
in
the
northern
hemisphere,
it
has
been
shown
that
concentrations
in
Arctic
have
declined
over
time
(
Li
and
Bidleman
2003
and
Shen
et
al.,
2005).
Data
collected
from
the
Canadian
and
Norwegian
Arctic
and
from
21
the
Bering
and
Chukchi
Seas
indicate
that
atmospheric
concentrations
of
 ­
HCH
have
declined
between
1973
and
1993.
It
was
concluded
that
these
data
are
reflective
of
decreased
use
of
technical
HCH
products,
especially
in
Asian
countries
(
Iwata
et
al.,
1993).
However,
in
ocean
water,
the
concentrations
of
the
 ­
HCH
isomer
have
remained
relatively
the
same
since
the
early
1980s
(
Li
et
al.,
1998).

Removal
of
foliar
and
broadcast
type
applications
and
uses
in
favor
of
low
rate
seed
treatments
will
most
likely
limit
the
amount
of
lindane
available
for
release
into
the
environment.
A
seed
leaching
study
indicates
that
the
lindane
residues
on
the
coated
seeds
were
loosely
bound
and
were
readily
released.
About
88%
of
lindane
from
treated
seeds
was
released
at
pH
7.0
solution
(
US
EPA,
2005).
A
field
study
conducted
by
Waite
et
al.
(
2001)
in
Saskatchewan,
Canada
demonstrated
volatilization
of
lindane
from
fields
planted
with
lindane­
treated
canola
seed.
They
reported
that
significant
quantities
(
12­
30%)
of
applied
lindane
volatilize
from
treated
canola
seed
to
the
atmosphere
during
the
growing
seasons
and
have
direct
implications
on
regional
atmospheric
concentrations
of
lindane.
They
have
also
estimated
that
a
range
of
66.4
to
188.8
tons
of
atmospheric
load
of
 ­
HCH
occurred
during
1997
and
1998
following
the
planting
of
canola
in
the
region
of
the
Canadian
prairies.
Poissant
and
Koprivnjak
(
1996)
reported
that
90%
of
elevated
 ­
HCH
concentration
in
the
atmosphere
at
Villeroy,
Quebec
in
1992
was
from
secondary
emissions
of
applied
lindane­
treated
corn,
while
the
rest
was
from
the
volatilization
of
residual
lindane
from
the
previous
year
seed
treatment.
Jianmin
et
al.
(
2003)
modeled
lindane
transport
and
deposition
to
the
Great
Lakes
from
usage
areas
in
the
Canada
prairies
and
corn
belt
regions
of
southern
Ontario
and
Quebec.
Results
showed
that
lindane
transport
to
the
Great
Lakes
during
spring­
summer
came
mainly
from
application
sites
in
the
prairies,
with
minor
contribution
from
the
corn
belt.
They
compared
the
modeled
concentration
with
the
monitoring
data
of
the
Integrated
Atmospheric
Deposition
Network
(
IADN)
sites,
which
were
within
50
to
134%
of
those
measured
during
summer,
16­
51%
in
fall
and
3­
20%
in
winter.

Recently,
seasonal
air
concentrations
of
HCH
isomers
were
monitored
using
Passive
Air
Samplers
(
PAS)
along
an
urban
to
rural
transect
in
Toronto,
Canada
(
Motelay­
Massei
et
al.,
2005).
Air
concentration
of
  
HCH
showed
little
spatial
and
temporal
variability,
which
ranged
from
36.9
pg/
m3
to
98.5
pg/
M3.
These
concentrations
are
consistent
with
the
levels
reported
for
IADN
sites
(
U.
S.
EPA,
2000)
and
background
level
for
North
America.
However,
the
air
concentration
of
 ­
HCH
showed
greater
variability.
The
air
concentrations
of
 ­
HCH
were
159
pg/
M3
to
1020
pg/
M3
in
the
rural
sites
during
the
spring­
summer
monitoring
period.
Similar
trend
of
air
concentrations
of
 ­
and
 ­
HCH
were
also
observed
by
Hoff
et
al.
(
1992)
in
Ontario,
Canada.
Both
studies
concluded
that
the
continuing
use
of
lindane
during
spring
is
likely
associated
with
higher
concentration
of
 ­
HCH
in
the
air
samples.
Analysis
of
1990
to
2001
data
from
the
IADN
also
confirmed
that
annual
agricultural
application
was
a
key
variable
in
explaining
the
annual
cycle
of
atmospheric
lindane
concentrations
(
Buehler
et
al.,
2004).

A
continent
wide
analysis
of
HCH
isomers
in
ambient
air
was
monitored
using
PAS
at
40
stations
across
North
America
(
Shen
et
al.,
2004).
The
impact
of
lindane
22
usage
was
shown
by
lower
 /
 
ratios
throughout
the
continent
interior
and
in
Mexico.
Highest
ratios
of
 /
 
HCHs
were
found
in
the
coastal
regions
of
Canada,
consistent
with
long­
range
atmospheric
transport
of
technical
HCH
and
/
or
volatilization
of
  
HCH
from
coastal
seawater.
Temporal
sampling
of
precipitation
from
Great
Lake
IADN
stations
from
1997
to
2002
showed
no
significant
decrease
of
 
and
 ­
HCHs,
but
significant
increases
in
 ­
HCH
at
all
stations
(
Carlson
and
Hites,
2004.)

B.
Ecological
Effects
1.
Endocrine
Disruption
Based
on
available
scientific
literature,
lindane
has
characteristics
of
an
endocrine
disrupting
compound.
There
are
data
to
support
this
contention
for
invertebrates,
birds,
fish,
amphibians
and
especially
in
mammals.
Mammalian
studies
have
shown
adverse
endocrine
effects
(
Raizada
et
al.
1980;
Uphouse
1987;
Cooper
et
al.
1989)
and
lindane
has
been
reported
to
disturb
male
mammalian
reproductive
functioning
(
Chowdhury
et
al.
1987;
Chowdhury
and
Gautam
1994;
Dalsenter
et
al.
1997;
Dalsenter
et
al.
1996;
Beard
and
Rawlings
1998).

Lindane
is
also
known
to
accumulate
in
fat
tissues
and
to
be
slowly
eliminated
in
milk
during
lactation
(
Pompa
et
al.
1994).
Milk
is
known
to
be
a
major
route
of
elimination
for
lipophilic
persistent
substances
stored
in
adipose
tissue.
The
milk:
plasma
concentration
ratio
for
lindane
indicates
a
much
more
efficient
excretion
of
the
compound
in
milk
(
Dalsenter
et
al.
1997).
Milk
possesses
a
great
affinity
on
liposoluble
substances
due
to
its
high
fat
content.
The
presence
of
lindane
in
mammalian
milk
exposes
nursing
offspring
during
critical
periods
of
post­
natal
development
(
Dalsenter
et
al.
1997).

Exposure
of
mammalian
neonates
to
lindane
during
lactation
induces
reproductive
hazards
to
male
offspring
rats
which
are
detectable
in
adulthood.
Dalsenter
et
al.
(
1997)
indicated
that
treatment
of
female
rats
on
day
15
of
pregnancy
with
only
a
single
dose
(
30
mg
lindane/
kg
of
body
weight)
affects
the
sexual
behavior
of
adult
male
offspring
by
altering
libido
and
by
reducing
testosterone
concentration
without
compromising
fertility.
Effects
to
offspring
may
be
due
to
the
indirect
interference
of
lindane
on
hormonal
regulation
in
males.
Pertubation
of
the
endocrine
system
during
early
stages
of
development
can
be
influenced
by
small
changes
of
hormonal
imbalance.
Oral
exposure
of
rats
to
lindane
for
5
days
during
gestation
caused
effects
in
adult
male
offspring
that
included
testicular
alterations,
reduced
sperm
head
counts
and
increased
chromatin
abnormalities
in
epididymal
sperm
(
Traina
et
al.
2003).

Oral
exposure
to
 ­
HCH
or
to
technical
grade
HCH
caused
degenerative
changes
in
male
reproductive
tissues
and
produced
sperm
abnormalities
in
rats
and
mice
(
Dikshith
et
al.
1991a;
Gautam
et
al.
1989;
Nigam
et
al.
1979;
Pius
et
al.
1990;
Chowdhury
and
Gautam
1990;
VanVelsen
et
al.
1986).
Oral
exposure
to
 ­
HCH
produced
findings
indicative
of
antiestrogenic
activity.
Effects
included
reduced
embryo
implantation
in
mice
(
Sicar
and
Lahiri
1989),
reduced
ovulation
rate
in
rabbits
(
Lindenau
et
al.
1994)
and
delayed
vaginal
opening,
disrupted
estrous
cycling,
and
reduced
uterine
weights
in
rats
23
(
Chadwick
et
al.
1988).
Oral
exposure
to
lindane
for
5
days
during
gestation
caused
effects
in
adult
male
offspring.

Exposure
of
mink
(
Mustela
vision)
to
dietary
gamma
HCH
resulted
in
a
60%
decrease
in
reproductive
success
and
reduced
testis
size
in
gamma
HCH
treated
third
generation
males
(
Beard
and
Rawlings
1998).

In
a
submitted
avian
reproduction
study
using
the
mallard
duck
(
MRID
448671­
01),
thyroid
weights
for
males
in
the
135
ppm
test
concentration
were
significantly
higher
than
those
measured
in
the
control.
Histopathology
revealed
microscopic
lesions
in
the
thyroid
glands
consisting
of
thyroid
follicular
distension
and
coalescence,
follicular
hypertrophy
and
follicular
hyperplasia.
These
lesions
were
more
apparent
at
the
135
ppm
than
at
45
ppm.
Results
of
analysis
of
the
gonads
of
either
sex
were
unremarkable
with
the
exception
of
the
possibility
of
reduced
spermatogenesis
in
the
group
receiving
45
ppm.

Exposure
to
lindane
(
0.1
ppb)
produced
statistically
significant
low
dose
sex
ratio
effects
(
71%
males)
in
wood
frogs,
suggesting
possible
hormesis
(
low
dose
affected
but
no
effects
in
upper
doses)
(
Serben
2003).
Low
dose
circulating
hormone
concentrations
were
also
significantly
higher
than
controls
(
corticosterone
increased
44%
and
T4:
T3
ratio
increased
30%).

In
addition,
Petit
et
al.
(
1997)
found
that
lindane
exhibited
estrogenic
activity
in
two
in
vitro
aquatic
bioassays.
Other
in
vitro
effects
suggesting
possible
endocrine
disruption
included
altered
sperm
responsiveness
to
progesterone
(
Silvestroni
and
Palleschi
1999)
and
inhibition
of
testicular
steroidogenesis
in
rat
Leydig
cells
(
Ronco
et
al.
2001).
In
another
in
vitro
study,
gamma
HCH
was
able
to
induce
the
expression
of
both
vitellogenin
and
estrogen
receptors
in
primary
culture
of
rainbow
trout
hepatocyte
further
demonstrating
its
estrogenic
activity
(
Fluoriot
et
al.,
1995).

And
finally,
exposure
to
environmental
concentrations
of
gamma
HCH
for
28
days
lead
to
significant
high
levels
of
oestradiol
and
low
levels
of
testosterone
in
the
hemolymph
of
all
gamma
HCH
treated
male
green
neon
shrimp
(
Neocaridina
denticulate)
(
Huang
et
al.,
2004).

2.
Toxicity
and
Effects
Lindane
is
moderately
toxic
to
birds
and
mammals
on
an
acute
exposure
basis.
Chronic
reproductive
effects
include
significant
reductions
in
egg
production,
growth
and
survival
parameters
in
birds,
and
decreased
body
weight
gain
in
mammals.
Lindane
is
highly
toxic
to
bees
on
an
acute
contact
basis
(
LD50s
ranged
from
0.20
to
0.56
:
g/
bee).

Lindane
exhibits
high
to
very
high
acute
toxicity
to
freshwater
fish
(
LC50
ranges
of
1.7
to
131
ppb)
and
freshwater
aquatic
invertebrates
(
LC50
ranges
of
10.0
to
520
ppb).
Chronic
effects
include
reduction
in
larval
growth
in
freshwater
fish
(
NOAEC=
2.9
:
g/
L)
and
decreased
reproduction
in
aquatic
invertebrates
(
NOAEC=
54
:
g/
L).
Also,
lindane
24
exhibits
high
to
very
high
acute
toxicity
to
estuarine/
marine
fish
and
ranges
from
moderately
to
very
highly
toxic
to
estuarine/
marine
aquatic
invertebrates.

Reproductive
and
population
effects
in
other
species
of
invertebrates
have
also
been
suggested.
Blockwell
et
al.
(
1999)
found
that
populations
of
H.
azteca
(
a
detritivorous
crustacean)
exposed
to
(
LOAEL=
13.5
ug
lindane/
L;
NOAEC=
6.9
ug
lindane/
L)
lindane
were
significantly
(
ANOVA,
p
<
0.001;
Tukey­
Kramer,
p
<
0.05)
smaller
than
control
populations
in
a
35
day
chronic
study.
Reduction
in
population
growth
was
observed
and
resulted
from
a
combination
of
toxicant
effects:
disruption
of
the
reproductive
behavior
patterns
of
adult
H.
azteca
and
a
reduction
in
the
growth
of
recruited
individuals
and
consequently
their
delayed
sexual
development.
This
value
is
similar
to
the
LOAECs
produced
from
other
chronic
lindane
toxicity
studies
conducted
with
freshwater
crustaceans:
19
:
g/
L
for
Daphnia
magna
in
a
64­
day
study
and
8.6
:
g/
L
in
a
17­
week
study
conducted
with
Gammarus
fasciatus
based
on
survivorship
and
reproductive
success
(
Macek
et
al.,
1976).
Furthermore,
an
LOAEC
of
9.9
ug
lindane/
L
was
generated
in
a
life
cycle
study
conducted
using
Chironomous
riparius
(
Insecta)
(
Taylor
et
al.
1993).
Lindane
has
also
previously
been
reported
to
reduce
juvenile
growth
of
the
European
amphipod
Gammarus
pulex
(
L.)
at
6.1
:
g/
L
in
a
14­
d
study
(
Blockwell
et
al.
1996).
However,
data
shows
that
concentrations
of
lindane
above
2.5
:
g/
L
(
found
in
Lake
Michigan
tributary
stream)
were
not
reported
as
occurring
in
any
aquatic
system
tested
(
ATSDR
1997).

Neurological
and
behavioral
alterations
are
principal
toxic
effects
of
lindane
in
animals
(
Hulth
et
al.
1976;
Joy
1982).
Chakravarty
et
al.
(
1986)
and
Chakravarty
and
Lahiri
(
1986)
found
that
when
domestic
ducks
were
force
fed
lindane
(
20
mg/
kg
of
body
weight
for
8
wks),
significant
egg­
shell
thinning,
reduced
clutch
size,
and
reduced
laying
frequencies
were
observed.
They
suggested
that
lindane
induced
estradiol
insufficiency
which
causes
inhibition
of
hepatic
ribonucleic
acid
(
RNA)
and
yolk
protein
synthesis,
thereby
preventing
transformation
of
moderately
differentiated
oocytes
to
mature
vitellogenic
follicles,
delaying
ovulation
and
thus
drastically
reducing
clutch
size.
Hoffman
and
Eastin
(
1982)
found
that
lindane
was
teratogenic
to
mallard
ducks
only
at
doses
that
were
greater
than
five
times
the
field
level
of
application,
but
did
find
that
lindane
was
much
more
toxic
on
a
lbs
per
acre
basis
when
administered
in
oil.
However,
lindane
in
the
diet
of
laying
hens
at
100
ppm
caused
reduced
hatchability
(
Whitehead
et
al.
1972)
and
at
25
ppm
the
same
effect
was
noted
in
Japanese
quail
(
Dewitt
and
George
1957).

3.
Endangered
Species
Many
taxonomic
groups
are
potentially
at
risk
from
exposure
to
lindane
treated
seed.
Animals
may
ingest
seed,
may
be
exposed
through
contaminated
soil
and
water,
and
through
atmospheric
transport.

The
Endangered
Species
Act
defines
the
action
area
for
a
Federal
action
as
being
the
footprint
of
possible
effects
stemming
from
the
action,
not
necessarily
limited
to
where
the
immediate
action
occurs.
For
screening­
level
purposes,
the
risk
assessment
25
conservatively
assumes
that
listed
species
are
co­
located
with
the
pesticide
treatment
area.
This
means
that
terrestrial
plants
and
wildlife
are
assumed
to
be
located
on
or
adjacent
to
the
treated
field
and
aquatic
organisms
are
assumed
to
be
located
in
a
surface
water
body
adjacent
to
the
treated
field.
This
assumption
places
the
listed
species
within
an
assumed
area
of
high
potential
exposure
to
the
pesticide.
Also,
since
lindane
is
transported
into
arctic
regions,
species
in
those
areas
may
also
be
exposed.

The
Agency
can
use
dose
response
relationships
(
slopes)
from
toxicity
studies
to
estimate
the
probability
of
acute
effects
associated
with
levels
of
concern.
This
information
serves
as
a
guide
to
establish
the
need
for
and
extent
of
additional
analysis
that
may
be
performed
using
Services­
provided
"
species
profiles"
as
well
as
evaluations
of
the
geographical
and
temporal
nature
of
the
exposure
to
ascertain
if
a
"
not
likely
to
adversely
affect"
determination
can
be
made.
The
degree
to
which
additional
analyses
are
performed
is
commensurate
with
the
predicted
probability
of
adverse
effects
from
the
comparison
of
the
dose
response
information
with
the
estimated
environmental
concentrations.
The
greater
the
probability
that
exposures
will
produce
effects
on
a
taxa,
the
greater
the
concern
for
potential
indirect
effects
for
listed
species
dependent
upon
that
taxa,
and
therefore,
the
more
intensive
the
analysis
on
the
potential
listed
species
of
concern,
their
locations
relative
to
the
use
site,
and
information
regarding
the
use
scenario
(
e.
g.,
timing,
frequency,
and
geographical
extent
of
pesticide
application).

The
probit
slope
response
relationship
is
evaluated
to
calculate
the
chance
of
an
individual
event
corresponding
to
the
listed
species
acute
LOCs.
If
information
is
unavailable
to
estimate
a
slope
for
a
particular
study,
a
default
slope
assumption
of
4.5
is
used
as
per
original
Agency
assumptions
of
typical
slope
cited
in
Urban
and
Cook
(
1986).

Terrestrial
Species
Analysis
of
raw
data
from
avian
and
mammalian
terrestrial
acute
toxicity
studies
for
lindane
estimate
slopes
of
2.456
for
birds
and
for
mammals,
a
default
value
of
4.5
was
used.
Based
on
these
slopes,
the
corresponding
estimate
chance
of
individual
mortality
following
exposure
is
1
in
1.42
x
102
for
birds
and
1
in
2.94
x
105
for
small
mammals.

Aquatic
Species
Analysis
of
raw
data
from
the
aquatic
acute
toxicity
studies
for
lindane
estimate
slopes
of
4.67
for
freshwater
fish,
a
default
value
of
4.5
for
freshwater
invertebrates,
a
default
value
of
4.5
for
estuarine/
marine
fish
and
3.07
for
estuarine/
marine
invertebrates.
Based
on
these
slopes,
the
corresponding
estimated
chance
of
individual
mortality
following
exposure
is
1
in
1.49
x
109
for
freshwater
fish,
1
in
4.17
x
108
for
freshwater
invertebrates,
1
in
4.17
x
108
for
estuarine/
marine
fish
and
1
in
3.08
x
104
for
estuarine/
marine
invertebrates.
26
IV.
HCH
Human
Health
Risk
Assessment
As
part
of
EPA's
broader
assessment
of
technical
HCH
and
its
isomers,
the
Agency
has
conducted
an
analysis
of
potential
health
effects
associated
with
exposure
to
these
other
forms
of
HCH.
Like
lindane,
these
HCH
isomers
exhibit
relatively
high
stability
and
their
global
production
for
many
years
has
lead
to
their
continued
presence
in
air,
soil,
surface
water,
ground
water,
and
drinking
water
as
discussed
in
Section
III
above.

This
analysis
summarizes
the
Agency's
previous
assessment
of
lindane
and
provides
a
review
of
available
toxicity,
health
effects,
and
exposure
data
on
the
byproduct
HCH­
isomers.
Because
the
most
predominant
byproducts
of
the
lindane
production
process
are
the
"­
and
$­
HCH
isomers
and
these
isomers
are
found
most
often
and
at
highest
concentrations
in
the
environment,
this
assessment
focuses
primarily
on
the
"­
and
$­
HCH
isomers.
As
part
of
this
assessment,
the
Agency
has
used
available
toxicity
and
exposure
data
to
conduct
a
preliminary
analysis
of
estimated
dietary
exposure
and
risk
to
indigenous
people
of
Alaska
from
the
"­
and
$­
HCH
isomers.
Information
for
this
human
health
assessment
of
"­
and
$­
HCH
isomers
has
been
drawn
from
a
variety
of
sources
including
published
reports
of
US
and
International
Governmental
Organizations
and
peer­
reviewed
literature
and
abstracts.

A.
Gamma­
HCH
Isomer
(
Lindane)

The
Agency
conducted
a
human
health
risk
assessment
to
support
the
2002
RED
for
lindane.
EPA's
assessment
included
estimates
of
risk
from
dietary,
drinking
water,
and
worker
exposures
to
lindane
resulting
from
its
use
as
a
seed
treatment.
Based
on
EPA's
assessment
of
dietary
and
drinking
water
risk
associated
with
use
of
lindane
as
a
pre­
plant
seed
treatment,
both
acute
and
chronic
aggregate
dietary
and
drinking
water
risks
from
lindane
are
below
EPA's
level
of
concern.
Because
lindane
persists
in
the
environment
and
has
long­
range
atmospheric
transport
potential,
the
Agency
also
performed
an
assessment
of
dietary
risk
resulting
from
subsistence
diets
of
indigenous
peoples
of
the
Arctic
region
of
the
U.
S.
(
Alaska)
who
rely
heavily
on
game
for
their
food
source.
That
analysis
indicated
that
dietary
risks
from
lindane
to
indigenous
people
of
Alaska
are
generally
not
of
concern.
EPA
also
assessed
risk
from
occupational
exposure
to
lindane
which
occurs
either
on­
farm
or
at
commercial
seed
treatment
facilities
to
workers
who
mix,
load
and
apply
lindane
as
a
seed
treatment,
and
workers
who
handle
or
plant
treated
seed.
Based
on
the
Agency's
assessment,
on­
farm
handling
of
the
lindane
dust
formulation
to
mix/
load
and
plant
treated
seed
resulted
in
risks
of
concern.
However,
required
mitigation
including
use
of
a
lower
seed
planting
rate
and
additional
personal
protective
equipment
reduced
these
risks
to
below
EPA's
level
of
concern
for
on­
farm
handlers.
Estimates
of
risk
from
commercial
seed
treatment
were
below
EPA's
level
of
concern
with
no
risk
mitigation
required.
Also,
the
Agency's
assessment
indicated
no
risk
concerns
for
post­
application
exposures
to
agricultural
workers.
The
Lindane
Risk
Assessment
for
Reregistration
Eligibility
Document
(
RED)
can
be
found
at
www.
regulations.
gov/
27
The
Agency
initially
did
not
assess
exposures
to
other
HCH
isomers
as
part
of
its
Human
Health
Risk
Assessment
for
Lindane.
Significantly,
however,
a
number
of
environmental
and
other
organizations
representing
indigenous
people
submitted
comments
strongly
advocating
that
EPA
conduct
a
risk
assessment
of
exposure
to
other
HCH­
isomers,
particularly
$­
HCH,
as
a
result
of
lindane
use.

B.
Alpha­
and
Beta­
HCH
Isomers
1.
Exposure
Pathways
As
mentioned
above,
HCH
has
been
released
to
the
environment
over
many
years
during
its
production
and
use
as
a
pesticide.
Technical­
grade
HCH,
once
used
as
an
insecticide
in
the
U.
S.,
has
not
been
produced
or
used
in
the
U.
S.
for
over
20
years.
However,
byproducts
of
HCH
production,
"­,
$­,
and
*­
HCH
isomers,
can
still
be
found
in
soil
and
water
because
they
are
mobile
and
persist
in
the
environment
(
see
Section
III
of
this
document).
The
only
sources
of
direct
exposure
to
these
compounds
in
the
US
are
hazardous
waste
sites
at
which
technical­
grade
HCH
and
other
HCH
isomers
were
disposed.
Sources
of
direct
exposure
in
other
countries
include
facilities
at
which
lindane
is
still
being
produced,
abandoned
pesticide
plants,
and
hazardous
waste
sites.
Individual
HCH
isomers
partition
into
various
media
at
different
rates
depending
on
the
physical
and
chemical
properties
of
each
isomer.
Technical
HCH
itself
is
not
found
in
the
environment.

The
long­
term
stability
of
HCH
isomers
combined
with
long
range
continental
transport
through
the
movements
of
the
air
and
ocean
currents
has
led
to
continued
detection
of
HCH
isomers
in
air,
soil,
surface
water,
groundwater,
and
drinking
water.
Monitoring
data
suggests
that
the
general
population
is
exposed
to
HCH
isomers
through
the
inhalation
of
ambient
air
and
consumption
of
contaminated
food
and
drinking
water.

HCH
isomers
are
among
the
most
abundant
organochlorine
contaminants
in
the
Arctic
environment.
Like
other
persistent
organic
pollutants,
"­
and
$­
HCH
isomers
in
particular
accumulate
in
colder
climates
such
as
the
Arctic,
where
they
are
trapped
by
low
evaporation
rates.
These
HCH
isomers
have
a
high
lipid
solubility,
concentrating
in
the
fatty
tissue
of
animals.
Once
in
the
Arctic,
they
enter
the
food
chain,
bioaccumulating
and
concentrating
in
whales,
seals,
polar
bears,
fish
and
other
arctic
animals.
Consequently,
indigenous
peoples,
who
rely
heavily
on
animal
fats
and
protein
in
their
traditional
diets
are
likely
to
be
more
highly
exposed
to
these
HCH
isomers.

2.
Hazard
Identification
Hazard
identification
and
toxicity
information
for
this
analysis
was
obtained
from
various
U.
S.
and
international
sources.
The
primary
source
of
toxicological
information
for
this
assessment
is
the
Agency
for
Toxic
Substances
and
Disease
Registry's
(
ATSDR)
Toxicological
Profile
for
Alpha­,
Beta­,
Gamma­,
and
Delta­
Hexachlorocyclohexhane,
August,
2005.
ATSDR's
public
health
statement
provides
comprehensive
information
on
health
effects
associated
with
exposure
to
"­,
$­,
(­,
and
*­
HCH.
Information
on
the
28
distribution
of
"­
and
$­
HCH
in
various
populations
was
obtained
largely
from
the
Center
for
Disease
Control
and
Prevention
(
CDC's)
National
Report
on
Human
Exposure
to
Environmental
Chemicals
and
from
and
the
Arctic
Monitoring
and
Assessment
Program
(
AMAP).

a.
Health
Effects
This
section
provides
an
overview
of
the
toxicology
and
health
effects
of
"­
and
$­
HCH,
the
majority
of
which
are
excerpted
directly
from
the
2005
ATSDR
Toxicological
Profile
for
HCH.
It
should
be
noted
that,
compared
to
lindane
((­
HCH),
toxicological
data
for
"­
and
$­
HCH
is
extremely
limited.

Alpha­
HCH
Based
on
EPA's
literature
search,
no
acute
toxicity
studies
are
available
for
"­
HCH.
Studies
examining
the
toxicity
of
"­
HCH
via
inhalation
and
dermal
routes
of
exposure
are
also
not
available.
A
number
of
subchronic
and
chronic
oral
toxicity
studies
in
animals
have
been
conducted
for
"­
HCH.
Studies
on
developmental
and
reproductive
effects
of
"­
HCH
are
not
available.

Based
on
available
animal
studies,
effects
on
liver
and
kidney
appear
to
be
the
most
sensitive
endpoints
following
repeated
oral
exposure
to
"­
HCH.
Liver
effects
were
reported
in
animals
following
intermediate­
or
chronic­
duration
exposure
to
"­
HCH
in
the
diet.
Studies
of
intermediate
duration
exposure
(
3 
48
weeks)
have
reported
slight
liver
effects
(
hypertrophied
liver
cells)
or
increased
liver
weight
in
mice
exposed
to
18
mg/
kg/
day
of
"­
HCH
(
Ito
et
al.
1973).
These
studies
were
limited
by
either
a
small
sample
size
or
lack
of
statistical
analysis.
Liver
hypertrophy
was
observed
in
rats
fed
with
45
mg/
kg/
day
of
"­
HCH
in
the
diet
for
24
or
48
weeks
(
Ito
et
al.
1975).
Administration
of
1.8
mg/
kg/
day
"­
HCH
in
the
diet
to
rats
for
15
or
30
days
resulted
in
increases
in
hepatic
cytochrome
P­
450
content,
hepatic
lipid
peroxidation,
and
hepatic
microsomal
superoxide
production
(
Barros
et
al.
1991).
Hepatomegaly
was
reported
in
mice
exposed
to
90
mg/
kg/
day
in
the
diet
for
50
weeks
(
Tryphonas
and
Iverson
1983).
Long­
term
exposure
to
lower
doses
of
"­
HCH
was
reported
to
result
in
fatty
degeneration
and
focal
necrosis
in
rats
exposed
to
56 
64
mg/
kg/
day
for
107
weeks
(
Fitzhugh
et
al.
1950).

Kidney
damage
(
nephritis
and
basal
vacuolation)
was
reported
in
rats
fed
72 
80
mg
"­
HCH/
kg/
day
for
an
average
of
35.9
weeks;
these
effects
were
not
seen
in
rats
fed
5
mg/
kg/
day.
Poor
survival
was
noted
in
both
control
and
treated
animals.
(
Fitzhugh
et
al.
1950).

Significantly
decreased
body
weight
gain
has
been
seen
in
rats
treated
orally
with
800
ppm
"­
HCH
(
Fitzhugh
et
al.
1950).
Neurological
effects
have
not
been
reported
in
animals
treated
with
"­
HCH.
Muller
et
al.
(
1981)
reported
no
delay
in
tail
nerve
conduction
velocity
in
rats
fed
5.1,
54.2,
or
106.2
mg
"­
HCH/
kg/
day
for
30
days.
29
"­
HCH
appears
to
be
carcinogenic
in
mice
and
rats
following
subchronic
and/
or
chronic
exposure.
Hepatomas
and
hepatocellular
carcinomas
have
been
reported
in
a
number
of
strains
of
mice
exposed
to
13 
95
mg/
kg/
day
for
16 
36
weeks
(
Hanada
et
al.
1973;
Ito
et
al.
1973,
1976;
Nagasaki
et
al.
1975;
Tsukada
et
al.
1979).
Tryphonas
and
Iverson
(
1983),
however,
reported
no
evidence
of
a
carcinogenic
effect
in
male
mice
exposed
to
90
mg
"­
HCH/
kg/
day
in
the
diet
for
50
weeks.
No
evidence
of
liver
carcinogenicity
was
reported
in
Wistar
rats
exposed
to
45
or
90
mg
"­
HCH/
kg/
day
in
the
diet
for
24
or
48
weeks
(
Ito
et
al.
1975;
Nagasaki
et
al.
1975);
high
mortality
was
observed
in
both
the
treated
and
control
groups.
However
"­
HCH
may
be
carcinogenic
in
rats
after
long­
term
exposure.
Ito
et
al.
(
1975)
reported
an
increased
incidence
of
hepatocellular
carcinoma
in
male
rats
exposed
to
50
and
75
mg
"­
HCH/
kg/
day
in
the
diet
for
72
weeks.
A
study
of
enzyme­
altered
liver
foci
in
rats
treated
first
with
the
tumor
initiator
N­
nitrosomorpholine,
and
then
20
mg
"­
HCH/
kg/
day
in
food
for
49
weeks,
found
that
the
tumor
promoter
activity
of
"­
HCH
is
apparently
due
to
increased
cell
proliferation
caused
by
a
lowering
of
the
cell
death
(
apoptosis)
rate
(
Luebeck
et
al.
1995).
In
another
study
in
rats,
additional
administration
of
35
mg/
kg/
day
of
"­
HCH
in
the
diet
for
65
weeks
inhibited
the
induction
of
liver
tumors
by
0.07
mg/
kg/
day
of
aflatoxin
B1
(
Angsubhakorn
et
al.
1981).
Exposure
to
other
HCH
isomers
($­,
(­,
*­,
HCH)
alone
did
not
result
in
hepatocellular
carcinoma.
However,
when
other
HCH
isomers
were
mixed
with
"­
HCH,
hepatocellular
carcinoma
was
observed.
Results
of
this
mixed
isomer
study
suggest
that
"­
HCH
is
itself
a
hepatocellular
carcinogen
or
acts
synergistically
with
other
HCH
isomers.

In
a
study
of
women
from
India,
blood
levels
of
"­
HCH
(
and
other
isomers)
were
found
to
be
higher
in
women
with
breast
cancer
when
compared
to
healthy
women
without
the
disease
(
Mathur
et
al.
2002).
In
this
study,
135
breast
cancer
patients
and
50
females
without
cancer
filled
out
questionnaires
and
were
evaluated
for
their
body
burden
of
pesticides
through
blood
testing.
"­
HCH
blood
levels
were
significantly
higher
in
breast
cancer
patients,
41 
50
years
of
age,
compared
to
women
of
the
same
age
without
the
disease.
Other
organochlorine
pesticides,
including
DDT
and
its
metabolites,
were
also
present
in
the
blood.
EPA's
Integrated
Risk
Information
system
(
IRIS)
currently
lists
"­
HCH
as
a
probable
human
carcinogen
based
on
the
incidence
of
hepatic
nodules
and
hepatocellular
carcinomas
observed
in
male
mice
administered
"­
HCH
in
the
diet
(
Ito
et
al.
1973).

The
available
genotoxicity
data
indicate
that
"­
HCH
has
some
genotoxic
potential
but
the
evidence
for
this
is
not
conclusive.
Oral
exposure
to
"­
HCH
was
reported
to
result
in
mitotic
disturbances
including
an
increased
mitotic
rate
and
an
increased
frequency
of
polyploid
hepatic
cells
in
rats
(
Hitachi
et
al.
1975).
"­
HCH
produces
DNA
fragmentation
in
primary
cultures
of
rat
and
human
hepatocytes,
but
not
in
mouse
hepatocytes
(
Mattioli
et
al.
1996).
"­
HCH
was
reported
to
bind
to
calf
thymus
DNA
in
the
presence
of
metabolic
activation
(
Iverson
et
al.
1984).
"­
HCH
increased
the
mitotic
rate
and
frequency
of
polyploid
cells
in
rat
hepatocytes
(
Hitachi
et
al.
1975).
DNA
repair
induction
was
absent
in
hepatocytes
from
all
three
species.
"­
HCH
has
been
observed
to
bind
to
mouse
liver
DNA
at
a
low
rate
(
Iverson
et
al.
1984).
30
Table
3
below
provides
a
profile
of
available
toxicity
studies
on
"­
HCH.

Table
3.
Alpha­
HCH
Toxicity
Profile
Study
Type
Species
Results
Reference
Subchronic
15
d
ad
libitum
feeding
study
Wistar
Rat
LOAEL
=
1.8
mg/
kg/
day
(
Male)
based
on
hepatic
effects
­
increased
cytochrome
P­
450
level,
superoxide
dismutase,
catalase,
and
lipid
peroxidation
activities
Barros
et
al.
1991.

30
d
ad
libitum
feeding
study
Wistar
Rat
LOAEL
=
1.8
mg/
kg/
day
(
Male)
based
on
hepatic
effects
­
increased
cytochrome
P­
450
level,
superoxide
dismutase,
catalase,
NADPHcytochrome
P­
450
reductase
activities,
and
lipid
peroxidation
Barros
et
al.
1991.

24
wk
ad
libitum
feeding
study
dd
Mouse
LOAEL
=
18
mg/
kg/
day
(
Male)
based
on
hepatic
effects
­
centrilobular
hypertrophy
Ito
et
al.
1973.

50
wk
ad
libitum
feeding
study
HPB
Mouse
LOAEL
=
90
mg/
kg/
day
(
Male)
based
on
hepatic
effects
­
hyperplastic
nodules
Tryphonas
et
al.
1983.

30
d
ad
libitum
feeding
study
Wistar
Rat
NOAEL
=
106.2
mg/
kg/
day
Muller
et
al.
1981.

Chronic
107
wk
ad
libitum
feeding
study
Wistar
Rat
NOAEL
=
0.8
mg/
kg/
day
LOAEL
=
3.5
mg/
kg/
day
(
Female)
based
on
hepatic
effects
­
very
slight
to
slight
microscopic
damage
in
the
absence
of
gross
liver
damage;
32%
increased
liver
weight
Fitzhugh
et
al.
1950.

Cancer
20
wk
ad
libitum
feeding
study
Wistar
Rat
CEL
=
2
mg/
kg/
day
(
Female)
based
on
increase
in
preneoplastic
hepatic
foci
Schröter
et
al.
1987.

32
wk
ad
libitum
feeding
study
dd
Mouse
CEL
=
18
mg/
kg/
day
(
Male)
based
on
hepatoma
Hanada
et
al.
1973.

wk
ad
libitum
feeding
study
dd
Mouse
CEL
=
45
mg/
kg/
day
(
Male)
based
on
hepatocellular
carcinoma
Ito
et
al.
1973.

16­
36
wk
ad
libitum
feeding
study
DDY
Mouse
CEL
=
90
mg/
kg/
day
(
Male)
based
on
hepatocellular
carcinoma
Ito
et
al.
1976.

24
wk
ad
libitum
feeding
study
DDY,
ICR,
DBA/
2,
C57BL/
6,
C3H/
He
Mouse
CEL
=
90
mg/
kg/
day
(
Male)
based
on
hepatocellular
carcinoma
Nagasaki
et
al.
1975.
31
Table
3.
Alpha­
HCH
Toxicity
Profile
Study
Type
Species
Results
Reference
24
wk
ad
libitum
feeding
study
HPB
Mouse
CEL
=
90
mg/
kg/
day
(
Male)
based
on
hyperplastic
nodules
and
adenomas
in
liver
Tryphonas
et
al,
1983
16­
36
wk
ad
libitum
feeding
study
DD
Mouse
CEL
=
90
mg/
kg/
day
(
Male)
based
on
hepatoma
Tskada
et
al.
1979
16­
36
wk
ad
libitum
feeding
study
NS
Rat
CEL
=
75
mg/
kg/
day
(
Male)
based
on
hepatocellular
carcinoma
Ito
et
al.
1975
Genotoxicity
Test
System
­
Mammallian
Cells
in
vivo
Mouse
liver
Weakly
Positive
­
DNA
Binding
Iverson
F,
Ryan
JJ,
Lizotte
R,
et
al.
1984.

Test
System
­
Mammallian
Cells
in
vivo
Rat
liver
Positive
­
Mitotic
Disturbances
Hitachi
M,
Yamada
K,
Takayama
S.
1975.

Test
System
­
Mammallian
Cells
in
vitvo
Calf
­
Thymus
DNA
Weakly
Positive
­
DNA
binding
with
activation
Not
tested
without
activation
Iverson
F,
Ryan
JJ,
Lizotte
R,
et
al.
1984.

CEL
=
cancer
effect
level
 
the
lowest
dose
of
chemical
in
a
study,
or
group
of
studies,
that
produces
significant
increases
in
the
incidence
of
cancer
(
or
tumors)
between
the
exposed
population
and
its
appropriate
control;
d
=
day(
s);
Gd
=
gestation
day;
Ld
=
lactation
day;
LOAEL
=
lowest­
observed­
adverseeffect
level;
M
=
male;
mo
=
month(
s);
NOAEL
=
no­
observed­
adverse­
effect
level;
wk
=
week(
s)

Beta­
HCH
The
toxicity
data
base
for
$­
HCH
includes
studies
on
acute/
short­
term
toxicity
via
the
oral
route,
subchronic
and
chronic
oral
toxicity
studies,
and
a
limited
number
of
studies
on
reproductive
effects.
Studies
examining
the
toxicity
of
$­
HCH
via
inhalation
and
dermal
routes
of
exposure
are
not
available.

Short­,
intermediate­,
and
long­
term
exposure
to
ß­
HCH
in
the
diet
is
associated
with
liver
effects
in
animals.
Studies
of
intermediate
duration
exposure
(
3 
48
weeks)
have
reported
slight
liver
effects
or
increased
liver
weight
in
mice
exposed
to
45
mg/
kg/
day
of
ß­
HCH.
(
Ito
et
al.
1973).
These
studies
were
limited
by
either
a
small
sample
size
or
lack
of
statistical
analysis.
Significant
increases
in
liver
weight
and
in
the
levels
of
hepatic
cytochrome
P­
450,
triglycerides,
phospholipids,
and
cholesterol
were
observed
in
rats
administered
90
mg/
kg/
day
ß­
HCH
in
the
diet
for
2
weeks
(
Ikegami
et
al.
1991a,
1991b);
decreases
in
cytochrome
c
reductase
activity
were
also
reported.
A
dosedependent
increase
in
liver
weight
was
noted
in
rats
exposed
for
13
weeks
to
0.18 
4.5
mg/
kg/
day
of
ß­
HCH;
the
increase
was
significant
at
doses
of
>
1
mg/
kg/
day
(
Van
Velsen
et
al.
1986).
Liver
cell
hypertrophy
was
reported
in
rats
fed
25
or
50
mg/
kg/
day
of
ß­
HCH
in
the
diet
for
24
or
48
weeks
(
Ito
et
al.
1975).
In
mice,
exposure
to
45
mg/
kg/
day
for
24
weeks
resulted
in
liver
cell
hypertrophy
(
Ito
et
al.
1973),
and
exposure
to
54 
57
32
mg/
kg/
day
for
32
weeks
resulted
in
hepatic
foci
of
degeneration
(
Hanada
et
al.
1973).
Hepatomegaly
was
reported
in
mice
exposed
to
90
mg/
kg/
day
in
the
diet
for
50
weeks
(
Tryphonas
and
Iverson
1983).
Chronic
exposure
to
lower
doses
of
ß­
HCH
resulted
in
fatty
degeneration
and
necrosis
in
the
liver
of
mice
fed
56 
64
mg/
kg/
day
for
107
weeks
(
Fitzhugh
et
al.
1950),
and
Thorpe
and
Walker
(
1973)
reported
liver
cancer
in
mice
fed
34
mg/
kg/
day
for
26
months.

Renal
effects
have
also
been
noted
in
rats
exposed
to
ß­
HCH
in
the
diet.
Srinivasan
et
al.
(
1984)
reported
significantly
increased
excretion
of
glucose
in
urine
and
increased
excretion
of
creatinine
and
urea
as
well
as
hypertrophy
and
degeneration
of
the
renal
tubular
epithelia
in
rats
exposed
to
72
mg
ß­
HCH/
kg/
day
for
up
to
2
weeks.
Van
Velsen
et
al.
(
1986)
reported
significantly
increased
kidney
weights
in
female
rats
exposed
to
0.18
mg
ß­
HCH/
kg/
day
for
13
weeks;
males
did
not
show
a
significant
increase
until
they
were
exposed
to
a
dose
of
4.5
mg/
kg/
day.
At
22.5
mg/
kg/
day,
both
males
and
females
exhibited
renal
calcinosis
in
the
outer
medulla;
however,
the
female
controls
also
exhibited
calcinosis.
The
study
authors
noted
that
renal
calcinosis
is
common
in
female
rats
but
that
this
finding
was
of
significance
in
males
(
Van
Velsen
et
al.
1986).
Fitzhugh
et
al.
(
1950)
also
examined
the
renal
effects
of
exposure
to
ß­
HCH
in
rats
that
died
after
an
average
of
4.4
weeks
and
found
nephritis
and
basal
vacuolation;
poor
survival
due
to
unspecified
causes
was
reported
in
both
control
and
treated
animals.

Significantly
decreased
body
weight
gain
has
been
seen
in
rats
treated
orally
250
mg/
kg
ß­
HCH
(
Fitzhugh
et
al.
1950;
Van
Velsen
et
al.
1986).

Neurological
effects
have
been
reported
in
rats
exposed
to
ß­
HCH.
Mice
treated
with
57
or
190
mg/
kg/
day
ß­
HCH
for
30
days
developed
ataxia
within
1
week
of
treatment
(
Cornacoff
et
al
1988).
Muller
et
al.
(
1981)
reported
a
significant
delay
in
tail
nerve
conduction
velocity
in
rats
fed
66.3
mg
ß­
HCH/
kg/
day
for
30
days.
Van
Velsen
et
al.
(
1986)
reported
ataxia,
progressive
inactivity
and
coma
in
rats
exposed
to
25
mg
ß­
HCH/
kg/
day
for
2
weeks.

Decreased
lymphoproliferative
responses
to
mitogens
were
seen
in
mice
exposed
to
60
mg/
kg/
day
ß­
HCH
in
the
diet
for
30
days
(
Cornacoff
et
al.
1988).
There
were
no
associated
changes
in
immunoglobulins,
red
blood
cell
counts,
or
histology
of
the
thymus,
spleen,
or
lymph
nodes.
Cortical
atrophy
of
the
thymus
was
observed
in
rats
fed
22.5 
25
mg/
kg/
day
ß­
HCH
(
Van
Velsen
et
al.
1986).
Exposure
to
22.5
mg
ß­
HCH
/
kg/
day
in
the
diet
for
13
weeks
in
rats
resulted
in
a
statistically
significant
decrease
in
numbers
of
red
blood
cells
and
white
blood
cells
and
reduced
hemoglobin
and
packed
cell
volume
values
(
Van
Velsen
et
al.
1986).

Oral
exposure
or
rats
and
mice
to
ß­
HCH
has
resulted
in
degeneration
of
male
reproductive
organs
and
sperm
abnormalities.
Histological
effects
on
the
testes
and
uterus,
as
well
as
increases
in
sperm
abnormalities
and
decreases
in
sperm
counts,
have
been
observed
in
rats
orally
exposed
to
generally
high
doses
of
ß­
HCH
(
Dikshith
et
al.
1991;
Gautam
et
al.
1989;
Nigam
et
al.
1979;
Pius
et
al.
1990;
Roy
Chowdhury
and
33
Gautam
1990;
Van
Velsen
et
al.
1986).
Atrophy
of
the
ovaries
and
testes,
hyperplastic
and
vacuolized
endometrial
epithelium,
degeneration
of
the
seminiferous
tubules,
and
disruption
of
spermatogenesis
were
seen
in
rats
exposed
to
22.5 
25
mg
ß­
HCH/
kg/
day
in
the
diet
for
13
weeks
(
Van
Velsen
et
al.
1986).
In
a
reproductive
toxicity
study
in
rats,
dietary
exposure
to
20
mg/
kg/
day
of
ß­
HCH
during
gestation
caused
increased
fetal
deaths
within
5
days
of
birth
and
5
mg/
kg/
day
of
ß­
HCH
during
gestation
and
lactation
resulted
in
increased
liver
weights
of
pups
(
Srinivasan
et
al.
1991a).
In
a
reproductive
toxicity
study
in
mice,
oral
exposure
to
60
mg
ß­
HCH/
kg
for
30
days
resulted
in
normal
uteri
and
reproductive
cycling
in
female
mice
(
Cornacoff
et
al.
1988).

ß­
HCH
has
not
been
found
to
be
carcinogenic
in
Wistar
rats
exposed
to
25
or
50
mg/
kg/
day
in
the
diet
for
24
or
48
weeks
(
Ito
et
al.
1975)
or
in
dd
mice
exposed
to
18 
120
mg/
kg/
day
in
the
diet
for
24
or
32
weeks
(
Hanada
et
al.
1973;
Ito
et
al.
1973).
These
studies
were,
in
general,
of
short
duration,
used
a
small
number
of
animals,
or
failed
to
examine
all
of
the
animals.
Thorpe
and
Walker
(
1973)
reported
an
increased
incidence
of
hepatocellular
carcinomas
in
CF1
mice
exposed
to
26
mg/
kg/
day
in
the
diet
for
104
weeks.
This
is
the
only
chronic
study
evaluated
by
ATSDR
from
which
to
estimate
cancer
risk
from
exposure
to
ß­
HCH.
The
study
is
limited
by
the
use
of
only
one
nonzero
dose
group.
Also,
the
use
of
incidence
of
liver
tumors
alone
in
mice
to
predict
a
compound's
carcinogenicity
in
humans
may
be
equivocal
(
Vesselinovitch
and
Negri
1988).
A
diversity
of
factors
has
been
shown
to
influence
the
development
of
liver
cell
tumors
in
mice,
such
as
the
strain
of
the
mice
(
Nagasaki
et
al.
1975),
the
protein
or
calorific
value
of
the
diet
(
Tannenbaum
and
Silverstone
1949),
and
the
microbial
flora
of
the
animals
(
Roe
and
Grant
1970).

In
a
study
of
women
from
India
in
which
135
breast
cancer
patients
and
50
females
without
cancer
were
evaluated
for
their
body
burden
of
pesticides
through
blood
testing,
blood
levels
of
ß­
HCH
were
found
to
be
higher
in
women
with
breast
cancer
when
compared
to
healthy
women
without
the
disease.
ß­
HCH
blood
levels
were
significantly
higher
in
breast
cancer
patients,
31 
50
years
of
age,
compared
to
those
without
the
disease.
(
Mathur
et
al.
2002).
Other
organochlorine
pesticides,
including
DDT
and
its
metabolites,
were
also
present
in
the
blood.
In
contrast
to
the
Mathur
et
al.
(
2002)
findings,
there
were
no
positive
associations
between
serum
levels
of
ß­
HCH
and
incidence
of
breast
cancer
in
studies
of
95
Mexican
women
(
Lopez­
Carrillo
et
al.
2002),
150
Norwegian
women
(
Ward
et
al.
2000),
or
240
women
from
Denmark
(
Hoyer
et
al.
1998).
The
risk
for
endometrial
cancer
was
also
not
associated
with
ß­
HCH
serum
concentrations
in
a
study
of
90
women
from
the
United
States
(
Sturgeon
et
al.
1998).
Levels
of
ß­
HCH
in
surgically
removed
breast
tissue
samples
from
65
women
in
Germany
were
not
indicative
of
malignant
breast
disease;
there
was
no
significant
difference
between
the
levels
of
ß­
HCH
in
the
breast
tissue
surrounding
malignant
and
benign
breast
disease
(
Guttes
et
al.
1998).

IRIS
currently
lists
ß­
HCH
as
a
possible
human
carcinogen
based
on
the
incidence
of
hepatic
nodules
and
hepatocellular
carcinomas
observed
in
male
mice
administered
ß­
HCH
at
a
single
dose
level
in
the
diet
(
Thorpe
and
Walker
1973).
34
The
available
genotoxicity
data,
while
severely
limited,
indicate
that
ß­
HCH
has
some
genotoxic
potential
but
the
evidence
for
this
is
not
conclusive.
Positive
results
were
reported
in
bone
marrow
cells
of
rats
exposed
to
ß­
HCH
(
Shimazu
et
al.
1972).
ß­
HCH
has
been
observed
to
bind
to
mouse
liver
DNA
at
a
low
rate
(
Iverson
et
al.
1984).

Table
4
provides
a
profile
of
available
toxicity
studies
on
ß­
HCH.

Table
4.
Beta­
HCH
Toxicity
Profile
Study
Type
Species
Results
Reference
Acute
2
wk
ad
libitum
feeding
study
Sprague­
Dawley
Rat
LOAEL
=
90
mg/
kg/
day
(
Male)
based
on
hepatic
effects
­
increased
triglycerides,
phospholipids
and
cholesterol,
increased
cytochrome
C
reductase
and
decreased
glutathione
peroxidase
Increased
relative
liver
weight
and
cytochrome
P­
450
levels
and
decreased
hepatic
vitamin
A
levels
Ikegami
et
al.
1991a.

Ikegami
et
al.
1991b.

14
day
ad
libitum
feeding
study
Wistar
Rat
LOAEL
=
72
mg/
kg/
day
(
Male)
based
on
renal
effects
­
tubular
degeneration,
distention
of
glomeruli,
swelling
of
tubular
epithelia,
22%
increase
in
kidney
weight,
altered
excretion
patterns
Srinivasan
et
al.
1984.

1
wk
ad
libitum
feeding
study
B6C3F1
Mouse
NOAEL
=
19
mg/
kg/
day
(
Female)
LOAEL
=
57
mg/
kg/
day
based
on
neurological
effects
­
ataxia
lateral
recumbancy
observed
at
190
mg/
kg/
day
Cornacoff
et
al.
1988.

2
wk
ad
libitum
feeding
study
Wistar
Rat
NOAEL
=
4.5
mg/
kg/
day
(
Female)
5
mg/
kg/
day
(
Male)
LOAEL
=
22.5
mg/
kg/
day
(
Male)
25
mg/
kg/
day
(
Male)
based
on
ataxia,
inactivity
Van
Velsen
et
al.
1986.

Sub­
Chronic
13
wk
ad
libitum
feeding
study
Wistar
Rat
LOAEL
=
4.5
mg/
kg/
day
(
Male)
based
on
hepatic
effects
­
hyalinization
of
centrilobular
cells,
focal
cell
necrosis,
increased
mitoses
LOAEL
=
22.5
based
on
15%
decreased
body
wt
(
Female);
ataxia,
coma
(
Male).
Van
Velsen
et
al.
1986.

32
wk
ad
libitum
feeding
study
dd
Mouse
NOAEL
=
20
mg/
kg/
day
LOAEL
=
54
mg/
kg/
day
(
Male)
based
on
hepatic
effects
­
nuclear
irregularities
in
foci
of
enlarged
hepatocytes
Hanada
et
al.
1973.

24
wk
ad
libitum
feeding
study
dd
Mouse
LOAEL
=
45
mg/
kg/
day
(
Male)
based
on
hepatic
effects
­
centrilobular
hypertrophy
Ito
et
al.
1973.
35
Table
4.
Beta­
HCH
Toxicity
Profile
Study
Type
Species
Results
Reference
13
wk
ad
libitum
reproduction
feeding
study
Wistar
Rat
NOAEL
=
0.2
mg/
kg/
day
(
Female)
0.9
mg/
kg/
day
(
Male)
LOAEL
=
1
mg/
kg/
day(
Female)
increased
absolute
ovary
and
uterous
weights;
4
mg/
kg/
day
(
Male)
decreased
testes
weight;
Effects
observed
at
22.5/
25
mg/
kg/
day
include
ataxia,
coma,
atrophy
of
ovary
and
testes,
hyperplastic
and
vacuolized
endometrium
epithelium
in
uterus
Van
Velsen
et
al.
1986.

30
day
ad
libitum
feeding
study
B6C3F1
Mouse
NOAEL
=
20
mg/
kg/
day
LOAEL
=
60
mg/
kg/
day
(
Female)
based
on
decreased
lymphoproliferative
responses
to
T­
cell
mitogens,
decreased
natural
killer
cytolytic
activity
Cornacoff
et
al.
1988.

30
day
ad
libitum
feeding
study
Wistar
Rat
LOAEL
=
66.3
mg/
kg/
day
(
Male)
based
on
reduced
tail
nerve
conduction
velocity
Muller
et
al.
1981.

Reproductive
feeding
study
(
Gd
0­
21;
Ld
1­
28)
Wistar
Rat
LOAEL
=
5
mg/
kg/
day
based
on
increased
liver
weight
in
pups
exposed
during
gestation
and
lactation;
increased
pup
mortality
occurred
at
20
mg/
kg/
day
Srinivasan
et
al.
1991b.

Chronic
107
wk
ad
libitum
feeding
study
Wistar
Rat
LOAEL
=
0.8
mg/
kg/
day
(
Female)
based
on
very
slight
microscopic
damage
in
the
absence
of
gross
liver
damage,
33%
increase
in
liver
weight
Fitzhugh
et
al.
1950.

Cancer
20
wk
ad
libitum
feeding
study
Wistar
Rat
CEL
=
3
mg/
kg/
day
(
Female)
based
on
increase
in
preneoplastic
hepatic
foci
Schröter
et
al.
1987.

104
wk
ad
libitum
feeding
study
CF1
Mouse
CEL
=
34
mg/
kg/
day
(
Female)
based
on
hepatocellular
carcinoma
Thorpe
et
al.
1973.

Genotoxicity
Test
System
­
Mammallian
Cells
in
vivo
Rat
­
Bone
marrow
Positive
­
Chromosomal
Aberrations
Shimazu
et
al.
1972.

CEL
=
cancer
effect
level
 
the
lowest
dose
of
chemical
in
a
study,
or
group
of
studies,
that
produces
significant
increases
in
the
incidence
of
cancer
(
or
tumors)
between
the
exposed
population
and
its
appropriate
control.
d
=
day(
s);
Gd
=
gestation
day;
Ld
=
lactation
day;
LOAEL
=
lowest­
observedadverse
effect
level;
M
=
male;
mo
=
month(
s);
NOAEL
=
no­
observed­
adverse­
effect
level;
wk
=
week(
s)

b.
Metabolism,
Absorption
and
Distribution
This
section
provides
an
overview
of
the
available
toxicokinetic
data
on
"­
and
$­
HCH.
The
primary
sources
of
information
for
this
section
are
the
2005
Final
ATSDR
Toxicological
Profile
for
HCH,
the
CDC's
National
Report
on
Human
Exposure
to
Environmental
Chemicals,
and
the
Arctic
Monitoring
and
Assessment
Program
(
AMAP).
36
The
CDC's
National
Report
provides
an
ongoing
assessment
of
the
exposure
of
the
U.
S.
population
to
environmental
chemicals
using
biomonitoring.
The
CDC's
Third
National
Report
released
in
2005
presents
biomonitoring
exposure
data
for
the
U.
S.
population
over
the
period
1999
­
2002
for
organochlorine
pesticides,
including
HCH.
Chemicals
or
their
metabolites
were
measured
in
blood
and
urine
samples
from
a
random
sample
of
participants
from
the
National
Health
and
Nutrition
Examination
Survey
(
NHANES)
conducted
by
CDC's
National
Center
for
Health
Statistics.
NHANES
is
a
series
of
surveys
designed
to
collect
data
on
the
health
and
nutritional
status
of
the
U.
S.
population.
AMAP
was
established
in
1991
to
monitor
identified
pollution
risks
and
their
impact
on
Arctic
ecosystems.
Under
the
AMAP
Human
Health
Program
all
circumpolar
countries
in
1994
agreed
to
monitoring
of
specific
human
tissues
for
contaminants
in
the
Arctic
to
address
concerns
of
increased
human
exposure
in
the
Arctic
and
possible
effects
of
a
number
of
organochlorine
contaminants.
Canada
coordinated
and
funded
this
circumpolar
activity.

Alpha­
HCH
Information
on
absorption,
metabolism
and
distribution
of
"­
HCH
is
extremely
limited.
In
an
absorption
study
of
technical­
grade
HCH
administered
in
the
feed
for
14
days,
the
average
absorption
of
"­
HCH
was
97.4%
(
Albro
and
Thomas
1974).
The
major
urinary
metabolites
formed
in
rats,
following
intermediate
oral
exposure
to
"­
HCH,
were
identified
as
tri­
and
tetrachlorophenols.
In
the
brain
of
rats,
"­
HCH
has
been
found
to
accumulate
preferentially
in
the
white
matter,
an
area
containing
lipid­
rich
myelin,
as
opposed
to
gray
matter
(
Portig
et
al.
1989).
Preferential
accumulation
of
"­
HCH
in
fatty
tissues
was
observed
in
gavage.
The
overall
distribution
pattern
indicated
the
that
"­
HCH
preferentially
distributes
first
to
fat,
then
kidney,
liver,
brain,
and
blood
in
declining
order
(
Eichler
et
al.
1983).

Information
on
the
distribution
of
the
HCH
isomers,
following
inhalation
by
humans
is
available
from
studies
of
humans
exposed
to
HCH
in
the
workplace.
Air
concentrations
of
"­
HCH
(
0.002 
1.99
mg/
m3)
were
associated
with
concurrent
mean
blood
serum
levels
in
workers
of
69.6
:
g/
L
(
Baumann
et
al.
1980).
Siddiqui
et
al.
(
1981a)
found
adipose
levels
of
0.1 
1.5
ppm
of
"­
HCH
in
the
tissues
collected
during
an
autopsy
case
study
conducted
in
India.

Beta­
HCH
Information
on
absorption,
metabolism
and
distribution
of
ß­
HCH
is
more
prevalent
than
"­
HCH
but
is
also
relatively
limited.
In
an
absorption
study
of
technicalgrade
HCH
administered
in
the
feed
for
14
days,
the
average
absorption
of
ß­
HCH
was
90.7%
(
Albro
and
Thomas
1974).
The
major
urinary
metabolites
formed
in
rats,
following
intermediate
oral
exposure
to
ß­
HCH,
were
identified
as
tri­
and
tetrachlorophenols.
The
elimination
of
ß­
HCH
was
investigated
in
a
group
of
40
former
workers
of
a
(­
HCH­
producing
plant
by
analyzing
at
least
two
blood
specimens
from
different
time
points
between
1952
and
1980.
The
median
half­
life
of
ß­
HCH
was
7.2
years,
calculated
by
concentrations
in
whole
blood,
and
7.6
years,
calculated
by
37
concentrations
in
extractable
lipids
assuming
first
order
kinetics
for
excretion
(
Jung
et
al.
1997).
Information
on
the
distribution
of
ß­
HCH
is
available
from
studies
in
which
laboratory
animals
were
acutely
exposed
by
ingestion
(
Chand
and
Ramachandran
1980;
Eichler
et
al.
1983;
Srinivasan
and
Radhakrishnamurty
1983b).
These
studies,
which
examined
the
overall
distribution
pattern
of
(­
and
ß­
HCH,
indicate
that
ß­
HCH
is
primarily
stored
in
the
fat
of
rats
acutely
exposed
for
5,
10,
or
15
days.
The
distribution
pattern
for
ß­
HCH
was
found
to
be
in
the
following
order:
fat
>
kidney
>
lungs
>
liver
>
muscle
>
heart
>
spleen
>
brain
>
blood
(
Eichler
et
al.
1983).
There
is
also
evidence
that
ß­
HCH
accumulates
in
tissues
to
a
greater
degree
than
(­
HCH
except
in
the
brain
(
Srinivasan
and
Radhakrishnamurty
1983b).
This
accumulation
increases
with
increasing
dose
and
treatment
period
for
ß­
HCH.
The
greater
accumulation
of
ß­
HCH
in
tissues
is
expected
since
this
isomer
is
known
to
be
metabolized
more
slowly.

Information
on
the
distribution
of
the
HCH
isomers
following
inhalation
by
humans
is
available
from
studies
of
humans
exposed
to
HCH
in
the
workplace.
Air
concentrations
of
ß­
HCH
(
0.001 
0.38
mg/
m3)
were
associated
with
concurrent
mean
blood
serum
levels
in
workers
of
190.3
:
g/
L
(
Baumann
et
al.
1980).
Accumulation
of
ß­
HCH
has
been
shown
to
increase
approximately
linearly
with
time
of
exposure
(
Baumann
et
al.
1980).
Siddiqui
et
al.
(
1981a)
found
adipose
levels
of
0.06 
0.9
ppm
of
ß­
HCH
in
the
tissues
collected
during
an
autopsy
case
study
conducted
in
India.

The
CDC
measured
serum
$­
HCH
in
a
subsample
of
NHANES
participants
aged
12
years
and
older.
Participants
were
selected
within
the
specified
age
range
to
be
a
representative
sample
of
the
U.
S.
population.
Because
of
its
longer
half­
life,
$­
HCH
is
usually
the
isomer
with
the
highest
concentration
in
the
general
population.
The
95th
percentile
level
of
ß­
HCH
for
the
current
2001­
2002
NHANES
subsample
is
similar
to
the
corresponding
95th
percentile
values
reported
in
a
study
of
adults
in
Germany
(
Wilhelm
et
al.,
2003).
Another
study
of
New
Zealand
adults
older
than
age
15
years
that
recently
reported
ß­
HCH
mean
levels
to
be
19.7
ng/
gram
of
lipid
(
Bates
et
al.,
2004),
which
are
slightly
higher
than
mean
levels
reported
for
the
1999­
2000
subsample.
The
levels
of
ß­
HCH
in
serum
shown
in
the
CDC
Report
are
far
below
a
biological
tolerance
level
of
25
:
g/
L
(
approximately
4,200
ng/
gram
of
serum
lipid)
in
serum
or
plasma
for
workers
at
the
end
of
their
shifts
as
defined
by
the
Deutsche
Forschungsgemeinchaft
(
2000).

According
to
the
CDC
report,
ß­
HCH
levels
in
the
U.
S.
population
have
been
declining
since
1970
(
Radomski
et
al.,
1971;
Stehr­
Green
et
al.,
1989;
Kutz
et
al.,
1991;
Sturgeon
et
al.,
1998).
Kutz
et
al.
(
1991)
estimated
that
in
1970
nearly
100%
of
the
U.
S.
population
had
detectable
ß­
HCH
in
adipose
tissue
and
that
80%
had
detectable
concentrations
in
1980,
with
the
mean
adipose
ß­
HCH
level
decreasing
from
0.37
:
g/
gram
of
lipid
(
370
ng/
gram)
in
1971
to
0.10
:
g/
gram
of
lipid
(
100
ng/
gram)
in
1983.
In
1976,
the
median
serum
lipid­
adjusted
level
of
ß­
HCH
was
119
ng/
gram
for
a
control
population
of
7,712
Danish
females
(
Hoyer
et
al.,
1998).
The
large
difference
between
these
1976
levels
and
current
U.
S.
levels
may
represent
a
global
change
in
levels
over
time.
Age­
related
increases
in
the
levels
of
ß­
HCH
have
previously
been
observed
by
the
German
Commission
on
Biological
Monitoring
(
Ewers
et
al.,
1999).
In
addition,
such
an
38
age
relationship
was
observed
previously
in
both
a
nonrandom
subsample
from
the
NHANES
II
(
1976­
1980)
and
for
ß­
HCH
levels
in
adipose
tissue
(
Stehr­
Green
et
al.,
1989;
Kutz
et
al.,
1991).
Also,
higher
levels
in
females
than
in
males
had
been
observed
for
ß­
HCH
levels
in
serum
(
Stehr­
Green
et
al.,
1989),
but
not
in
adipose
tissue
(
Burns,
1974).
Table
5
provides
a
summary
of
95th
percentile
$­
HCH
serum
concentrations
for
the
U.
S.
population
aged
12
and
older
from
the
National
Health
and
Nutrition
Examination
Survey,
1999­
2002.

Table
5.
Beta­
HCH
­
95th
Percentile
Serum
Concentrations
Survey
Years
1999­
2000
and
2001­
02
Units
Total
Pop
12­
19
years
20+
years
Male
Female
Mexican
Americans
Non­
Hispanic
Blacks
Non­
Hispanic
Whites
Survey
Years
2001­
2002
ng/
g
of
lipid
49.3
8.44
46.2
29.2
54.5
84.4
45.9
33.5
ng/
g
of
serum
0.296
0.048
0.3212
0.200
0.386
0.612
0.226
0.220
Survey
Years
1999­
2000
ng/
g
of
lipid
68.9
11.4
73.4
44.6
81.1
139
48.9
51.3
ng/
g
of
serum
0.447
0.055
0.477
0.286
0.556
0.905
0.359
0.390
Monitoring
data
on
blood
levels
of
organochlorine
compounds
in
mothers
living
in
Arctic
or
circumpolar
countries
have
been
collected
through
AMAP.
Under
the
AMAP
Human
Health
Program,
seven
circumpolar
countries
initially
agreed
to
participate
in
the
AMAP
effort
to
monitor
specific
human
tissues
for
organochlorine
contaminants:
Denmark/
Greenland,
Canada,
Sweden,
Norway,
Russia,
Iceland
and
Finland.
The
United
States
(
Alaska)
later
joined
the
sampling
program.
Under
the
AMAP
protocol,
mothers
in
these
circumpolar
countries
contributed
blood
samples
for
14
PCB
congeners
and
13
organochlorine
pesticides,
including
ß
 
­
HCH.
During
pregnancy
or
at
child
birth,
samples
of
venous
blood
were
collected
from
pregnant
women
and
all
countries
were
able
to
complete
the
sampling
between
1994
and
1997.
The
number
of
mothers
in
each
country
ranged
from
143
in
Finland
to
20
in
the
United
States.
This
study
allowed
an
assessment
of
the
variation
of
contaminants
in
human
populations
around
the
circumpolar
north
using
common
sampling
and
analytical
protocols.
Results
of
the
study
were
recently
published
in
a
journal
article
(
Van
Oostdam
et
al.
2004).

Van
Oostdam
et
al.
reported
the
following
regarding
 
ß­
HCH
serum
concentrations.
ß­
HCH
is
markedly
higher
in
the
Russian
mothers
(
223
ug/
kg
lipid)
than
any
other
population
sampled
(
5.1
to
32
ug/
kg
lipid).
Icelandic
mothers
have
intermediate
levels
of
ß­
HCH,
which
are
more
than
70%
higher
than
Greenland
Inuit
mothers
and
300
to
600%
higher
than
all
other
circumpolar
groups.
Beta­
HCH
is
likely
elevated
in
the
Russian
mothers
due
to
the
use
of
HCH
for
the
control
of
agricultural
pests
or
other
insects.
The
source
of
ß
 
­
HCH
in
Iceland
is
not
so
obvious
as
the
use
of
lindane
(
gamma­
HCH)
was
banned
in
Iceland
in
1994,
technical
HCH
(
mixture
of
alpha,
beta
and
gamma­
HCH)
was
never
licensed
for
use
in
Iceland,
and
Icelandic
women
do
not
consume
large
amounts
of
marine
mammals
(
Olafsdottir,
2001).
Significantly
higher
levels
of
ß
 
­
HCH
are
seen
in
the
Kitikmeot
Inuit.
Variations
in
contaminant
levels
may
reflect
different
dietary
preferences
among
the
different
Canadian
Inuit
populations
plus
39
varying
contaminant
levels
in
animals
consumed
due
to
age
or
region,
differing
availability
of
country
foods
or
other
factors.
For
example,
some
populations
consume
more
seals
and
some
more
beluga
whales
and
the
variation
in
contaminant
level
based
on
food
species
and
location
has
been
well
documented
(
Kuhnlein
et
al.,
2000
and
Muir
et
al.,
1999).
These
differences
among
Inuit
populations
may
also
represent
different
deposition
patterns
associated
with
long­
range
transport
into
the
Arctic
from
different
source
locations.
Table
6
provides
a
summary
of
ß­
HCH
maternal
blood
serum
concentrations
in
the
Circumpolar
North
from
samples
collected
from
1994
to
1997
(
Van
Oostdam
et
al.
2004).

Table
6.
Beta­
HCH
Contaminants
in
Maternal
Blood
in
the
Circumpolar
North
(
ng/
g
lipid)
(
1994­
1997)
Inuit
Populations
Non­
Inuit
Populations
Greenland
Canada
Nunavaik
Canada
Kitikmeot
USA
Alaska
North
Slope
Norway
Sweden
Iceland
Russia
Finland
Geometric
mean
18
5.1
9.6
7.9
8.1
9.2
32
223
7.2
95th%
Confidence
Interval
15­
22
4.2­
6.1
8.1­
11
4.6­
13
6.8­
9.5
7.7­
11
26­
40
194­
225
1.6­
8.6
%
Detect
100
100
97
100
97
95
100
100
92
Van
Oostdam
et
al.
also
present
comparative
data
on
serum
concentrations
of
 
ß­
HCH
in
other
regions
of
the
world
as
shown
in
Table
7.

Table
7.
Beta­
HCH
Contaminants
in
Comparative
Populations
ng/
g
(
lipid)
Great
Lakes
Residents
Canada
Nukus
Uzbekistan
Catalonia
Spain
Veracruz
Mexico
Year(
s)
1992
NA
1997­
1999
1997­
1998
Mean
age
36
26
31
NA
N
48
18
72
90
GM1
or
median2
151
5932
1221
2512
Range
1.9­
37
350­
3287
1.1­
1247
ND­
857
3.
Child
Susceptibility
Information
on
the
specific
effects
resulting
from
"
and
ß­
HCH
exposure
in
children
is
extremely
limited.
Generally,
health
effects
observed
in
adults
should
also
be
of
potential
concern
in
children.
However,
children
may
be
more
susceptible
due
to
their
unique
physiology
and
behavior.
Vulnerability
often
depends
on
developmental
stage.
Potential
effects
on
offspring
resulting
from
parental
exposures
as
well
as
any
indirect
effects
on
the
fetus
and
neonate
resulting
from
maternal
exposure
during
gestation
and
lactation
are
of
particular
concern.
ß­
HCH
is
lipophilic
and
accumulates
in
maternal
adipose
tissue
and
may
be
mobilized
during
pregnancy
and
lactation.
Levels
of
ß­
HCH
in
placenta,
maternal
blood,
and
umbilical­
cord
blood
were
higher
in
cases
of
stillbirths
than
in
live­
born
cases;
however,
many
other
organochlorine
pesticides
were
present
that
40
could
have
contributed
to
stillbirths
(
Saxena
et
al.
1983).
(­,
"­,
*­,
and
total
HCH
maternal
blood
and
umbilical­
cord
blood
levels
were
also
higher
in
mothers
who
gave
birth
to
intra­
uterine
growth
retardation
(
IUGR)
babies
(
Siddiqui
et
al.
2003).

4.
Toxicological
Endpoints
of
Concern
for
Risk
Assessment
A
number
of
toxicological
organizations
world­
wide
have
conducted
doseresponse
assessments
to
identify
toxicological
endpoints
for
use
in
deriving
quantitative
estimates
of
risk
associated
with
exposure
to
alpha
and
beta
isomers
in
particular.
EPA
conducted
a
review
of
risk
values
from
US
and
international
health
organizations,
government
agencies
and
independent
groups
including
ATSDR,
Health
Canada,
International
Agency
for
Research
on
Cancer
(
IARC),
U.
S.
EPA,
the
Toxicology
Excellence
for
Risk
Assessment's
(
TERA's)
International
Toxicity
Estimates
for
Risk
database,
and
other
international
health
organizations.
This
section
provides
a
summary
of
available
risk
assessment
endpoints
for
"­
and
$­
HCH.

a.
Alpha­
HCH
Non­
Cancer
ATSDR
and
the
Netherlands
National
Institute
of
Public
Health
and
the
Environment
(
RIVM)
have
evaluated
the
non­
cancer
oral
toxicity
data
for
"­
HCH
and
derived
chronic
oral
risk
values
or
reference
doses
(
RfDs).
ATSDR
and
RIVM
also
evaluated
the
noncancer
inhalation
toxicity
data
for
"­
HCH.
ATSDR
concluded
that
no
inhalation
reference
concentrations
(
RfCs)
could
be
developed
for
isomers
of
"­
HCH
due
to
insufficient
data.
RIVM
derived
a
RfC
based
on
a
subchronic
inhalation
study
in
rats.

Chronic
Oral
RfD
ATSDR
derived
a
chronic
RfD
of
0.008
mg/
kg/
day
for
"­
HCH
based
on
a
NOAEL
of
0.8
mg/
kg/
day
for
liver
effects
in
rats
and
an
uncertainty
factor
(
UF)
of
100
(
10
for
extrapolation
from
animals
to
humans,
and
10
for
human
variability).

The
critical
NOAEL
was
identified
in
a
chronic
toxicity
study
in
which
groups
of
10
Wistar
rats
of
each
sex
were
exposed
to
a­
HCH
in
the
diet
for
up
to
107
weeks
at
estimated
doses
of
0,
0.7,
3.5,
7,
or
56
mg/
kg/
day
in
males
and
0,
0.8,
4,
8,
or
64
mg/
kg/
day
in
females
(
Fitzhugh
et
al.
1950).
End
points
included
clinical
signs,
body
weight,
food
consumption,
organ
weights,
gross
pathology,
and
histopathology.
No
exposure­
related
changes
occurred
at
the
low
dose
in
either
sex,
indicating
that
the
highest
NOAEL
is
0.8
mg/
kg/
day
in
females.
Liver
effects
were
qualitatively
described
in
both
sexes
at
higher
doses,
progressing
from
very
slight
histological
changes
with
no
gross
liver
pathology
at
3.5 
4
mg/
kg/
day,
slight
histological
changes
with
no
gross
pathology
at
7 
8
mg/
kg/
day,
and
moderate
histological
damage
accompanied
by
moderate
gross
pathology
at
56 
64
mg/
kg/
day.
The
hepatic
histopathological
changes
classified
as
moderate
included
hepatic
cell
atrophy,
fatty
degeneration,
and
focal
41
necrosis.
Non­
hepatic
effects
included
decreased
body
weight
gain,
slight
kidney
histopathology
(
focal
nephritis),
and
reduced
lifespan
at
56 
64
mg/
kg/
day.

RIVM
derived
a
chronic
tolerable
daily
intake
(
TDI)
or
RfD
of
0.001
mg/
kg­
day
based
on
a
NOAEL
of
0.1
mg/
kg­
day
for
hepato­
and
nephrotoxicity
observed
at
the
LOAEL
of
0.5
mg/
kg/
day
in
a
90­
day
oral
rat
study
(
Slooff
and
Matthijssen,
1988).
RIVM
applied
a
UF
of
100
(
10
each
for
extrapolation
from
animals
and
for
human
variability).
RIVM
did
not
consider
an
additional
UF
necessary
for
the
use
of
a
subchronic
study
because
chronic
studies
did
not
indicate
effects
at
dosages
lower
than
the
NOAEL/
LOAEL
observed
in
the
90­
day
study
(
Baars
et.
al.
RIVM
2001).

Inhalation
RfC
RIVM
derived
a
tolerable
concentration
in
air
(
TCA)
or
RfC
of
0.00025
mg/
m3
based
on
a
NOAEL
of
0.025
mg/
m3
for
liver
and
kidney
toxicity
observed
in
a
subchronic
inhalation
bioassay
in
rats
(
Slooff
and
Matthijsen,
1988).
RIVM
used
an
UF
of
100
(
10
each
for
inter­
and
intraspecies
extrapolation).
RIVM
did
not
consider
an
additional
UF
necessary
for
the
use
of
a
subchronic
study
because
chronic
studies
did
not
indicate
effects
at
dosages
lower
than
the
NOAEL/
LOAEL
observed
in
the
subchronic
study
(
Baars
et.
al.
RIVM
2001).

Cancer
ATSDR
and
U.
S.
EPA
have
evaluated
the
carcinogenicity
data
for
"­
HCH.
IRIS
classified
this
chemical
as
B2
­
probable
human
carcinogen
and
estimated
an
oral
slope
or
cancer
potency
factor
of
6.3
(
mg/
kg/
day)­
1
(
EPA
IRIS).
The
cancer
weight­
of­
evidence
classification
is
based
on
all
routes
of
exposure.

The
primary
basis
for
the
cancer
classification
and
slope
factor
is
that
dietary
"­
HCH
has
been
shown
to
cause
increased
incidences
of
liver
tumors
in
five
mouse
strains
and
in
Wistar
rats
(
Ito
et
al.,
1973a,
b,
1976;
Nagasaki
et
al.,
1972,
1975;
Hanada
et
al.,
1973;
Goto
et
al.,
1972;
Schulte­
Hermann
and
Parzefall,
1981).
Ito
et
al.
(
1973a)
treated
groups
of
20­
40
male
dd
mice
with
100,
250,
or
500
ppm
"­
HCH
in
the
diet
for
24
weeks.
They
observed
liver
nodules
and
hepatocellular
carcinomas
in
the
two
upper
dose
groups.
In
a
subsequent
study,
Ito
et
al.
(
1976)
maintained
male
DDY
mice
on
a
diet
containing
500
ppm
"­
HCH
for
16,
20,
24,
or
36
weeks.
This
was
followed
by
basal
diet
for
4,
8,
12,
16,
24,
or
36
weeks,
respectively.
Incidence
of
liver
tumors
increased
with
continuous
"­
HCH
administration.
Incidence
decreased,
however,
with
recovery
time.
At
24
weeks
most
lesions
observed
were
nodules,
but
by
60
or
72
weeks
the
tumors
were
primarily
hepatocellular
carcinomas.

Schulte­
Hermann
and
Parzefall
(
1981)
noted
an
increased
incidence
of
hepatic
nodules
and
hepatocellular
carcinomas
in
female
Wistar
rats
treated
with
approximately
20
mg/
kg/
day
"­
HCH
for
their
lifetime.
Male
Wistar
rats
(
18­
24
animals/
group)
were
fed
"­
HCH
in
the
diet
at
500,
1000,
or
1500
ppm
for
24,
48,
or
72
weeks.
Liver
nodules
and
carcinomas
were
observed
in
rats
fed
the
two
highest
doses
for
72
weeks.
Liver
42
nodules
only
developed
in
animals
fed
1000
ppm
for
48
weeks
(
Nagasaki
et
al.,
1975).
Nagasaki
et
al.
(
1972)
observed
liver
nodules
and
tumor
formation
in
male
dd
mice
fed
250
or
500
ppm
for
24
weeks,
but
not
in
those
consuming
100
ppm
"­
HCH.
Both
males
and
females
of
the
dd
strain
responded
in
a
dose­
dependent
fashion
with
liver
nodules
and
hepatomas
when
fed
100,
300,
or
600
ppm
dietary
"­
HCH
for
32
weeks,
followed
by
5­
6
weeks
basal
diet
(
Hanada
et
al.,
1973).
In
a
feeding
study
using
male
ICR­
JCL
mice,
"­
HCH
produced
hepatomas
in
100%
of
the
animals
(
Goto
et
al.,
1972).
Liver
tumors
have
been
observed
as
early
as
24­
26
weeks
(
Sugihara
et
al.,
1975).
Table
8
below
summarizes
the
established
toxicological
endpoints
for
alpha­
HCH.

Table
8.
Alpha­
HCH
Summary
of
Doses
and
Toxicological
Endpoints
Risk
Assessment
Exposure
Route
Dose/
Endpoint
Endpoint
Source
Chronic
Oral
NOAEL
=
0.8
mg/
kg­
day
UF
=
100
RfD
=
0.008
mg/
kg­
day
LOAEL
=
3.5­
4
mg/
kg­
day
based
on
liver
effects
observed
in
a
chronic
toxicity
study
in
rats
Fitzhugh
et
al.
1950
(
ATSDR,
2005)

Chronic
Oral
NOAEL
=
0.1
mg/
kg­
day
UF
=
100
RfD
=
0.001
mg/
kg­
day*
LOAEL
=
0.5
mg/
kgday
based
on
leucocytopenia
and
liver
changes
observed
in
a
subchronic
toxicity
study
in
rats
Slooff
and
Matthijsen,
1988
(
RIVM
2001)

Inhalation
NOAEL=
of
0.025
mg/
m3
UF
=
100
RfC
=
0.00025
mg/
m3
LOAEL
=
0.1
mg/
kgday
based
on
liver
and
kidney
toxicity
observed
in
a
subchronic
inhalation
bioassay
in
rats
Vermeire
et.
al.
1991
(
RIVM
2001)

Cancer
(
all
routes)
Classification
­
B2;
probable
human
carcinogen
(
EPA
IRIS
1993)
Q1*
=
6.3
(
mg/
kg/
day)­
1*
Basis
­­
Dietary
"­
HCH
has
been
shown
to
cause
increased
incidence
of
liver
tumors
in
five
mouse
strains
and
in
Wistar
rats.
(
Ito
et
al.,
1973a,
b,
1976;
Nagasaki
et
al.,
1972,
1975;
Hanada
et
al.,
1973;
Goto
et
al.,
1972;
Schulte­
Hermann
and
Parzefall,
1981).
*
Endpoint
used
in
dietary
risk
assessment
b.
Beta­
HCH
Non­
Cancer
ATSDR
evaluated
the
non­
cancer
oral
toxicity
data
for
$­
HCH
and
derived
acute
and
intermediate
term
or
subchronic
RfDs.
RIVM
derived
a
chronic
RfD
based
on
observations
of
infertility
in
a
reproduction
toxicity
assay
in
rats.
ATSDR
and
RIVM
also
evaluated
the
noncancer
inhalation
toxicity
data
for
$­
HCH
and
concluded
that
no
43
inhalation
reference
concentrations
(
RfCs)
could
be
developed
for
$­
HCH
due
to
insufficient
data.

Acute
Oral
RfD
ATSDR
derived
an
acute
oral
RfD
0.05
mg/
kg/
day
for
ß­
HCH
based
on
a
NOAEL
of
4.5
mg/
kg/
day
and
LOAEL
of
22.5
mg/
kg/
day
for
clinical
signs
of
ataxia
in
rats
(
Van
Velsen
et
al.
1986)
and
an
uncertainty
factor
of
100
(
10
forextrapolation
from
animals
to
humans
and
10
for
human
variability).

The
principal
study,
Van
Velsen
et
al.
(
1986),
is
a
13­
week
toxicity
study
in
which
groups
of
10
Wistar
rats
of
each
sex
were
exposed
to
estimated
dietary
doses
of
0,
0.18,
0.9,
4.5,
or
22.5
mg/
kg/
day
in
males,
or
0,
0.2,
1.0,
5,
or
25
mg/
kg/
day
in
females.
At
week
2
of
the
study,
two
male
and
two
female
rats
receiving
the
highest
dose
(
22.5
and
25
mg/
kg/
day,
respectively)
exhibited
clinical
signs
of
ataxia
and
became
progressively
inactive.
Within
3
days
of
the
first
signs
of
ataxia,
the
animals
became
comatose
and
were
sacrificed.
The
investigators
did
not
report
adverse
clinical
signs
at
the
other
dose
levels;
thus,
the
4.5
mg/
kg/
day
(
in
males
and
5
mg/
kg/
day
in
females)
dose
is
considered
a
NOAEL.
Similar
neurotoxic
effects
were
observed
in
an
immunotoxicity
study
in
which
groups
of
six
female
B6C3F1
mice
were
exposed
to
ß­
HCH
in
the
diet
at
estimated
doses
of
0,
19,
57,
or
190
mg/
kg/
day
for
up
to
30
days
(
Cornacoff
et
al.
1988).
Mice
receiving
57
or
190
mg/
kg/
day
showed
signs
of
ataxia
within
the
first
week
of
exposure.
The
signs
resolved
in
a
few
days
in
the
57
mg/
kg/
day
group,
whereas
approximately
80%
of
the
190
mg/
kg/
day
mice
became
laterally
recumbent
and
moribund.
No
ataxia
or
other
signs
of
neurotoxicity
occurred
at
19
mg/
kg/
day.
Other
effects
in
this
study
included
immunological
alterations
at
57
mg/
kg/
day
(
e.
g.,
decreased
lymphoproliferative
responses
to
T­
cell
mitogens
and
decreased
natural
killer
cell
activity),
but
these
end
points
were
only
evaluated
after
30
days
and
are
therefore
not
considered
to
be
consequences
of
acute
duration
exposure.
Support
for
neurotoxicity
as
the
critical
effect
for
acute
oral
exposure
to
ß­
HCH
is
provided
by
the
Cornacoff
et
al.
(
1988)
study
reporting
ataxia
after
1
week
of
exposure
to
57
mg/
kg/
day
and
a
study
by
Muller
et
al.
(
1981)
reporting
a
significant
reduction
in
tail
nerve
motor
conduction
velocity
in
rats
exposed
to
66
mg/
kg/
day
ß­
HCH
for
30
days.

Chronic
Oral
RfD
ATSDR/
U.
S.
EPA
­
ATSDR
derived
an
intermediate­
duration
oral
RfD
of
0.0006
mg/
kg/
day
for
exposure
to
ß­
HCH
based
on
a
LOAEL
of
0.18
mg/
kg/
day
for
liver
effects
in
a
13
week
rat
study
and
an
UF
of
300
(
3
for
use
of
a
minimal
LOAEL,
and
10
for
extrapolation
from
animals
to
humans,
10
for
human
variability).
For
purposes
of
this
assessment,
EPA
has
used
results
of
the
subchronic
rat
study
to
derive
a
chronic
RfD
of
0.00006
mg/
kg/
day
for
ß­
HCH.
For
the
chronic
RfD,
EPA
added
an
UF
of
10
to
account
for
use
of
a
subchronic
study
to
assess
chronic
effects.
The
critical
LOAEL
was
identified
in
a
13­
week
subchronic
toxicity
study
in
which
groups
of
10
Wistar
rats
of
each
sex
were
exposed
to
estimated
dietary
doses
of
0,
0.18,
0.9,
4.5,
or
22.5
mg/
kg/
day
in
males,
or
0,
0.2,
1.0,
5,
or
25
mg/
kg/
day
in
females
(
Van
Velsen
et
al.
1986).
End
points
that
were
examined
include
body
weight,
food
consumption,
hematology,
blood
44
biochemistry,
organ
weights,
gross
pathology,
and
histopathology.
Hepatic
effects
were
observed
that
included
hyalinization
of
centrilobular
cells
in
males
at
=
0.18
mg/
kg/
day
and
females
at
25
mg/
kg/
day;
increased
absolute
and
relative
liver
weight
in
both
sexes
at
=
0.9
mg/
kg/
day
in
males
and
=
1.0
mg/
kg/
day
in
females;
periportal
fat
accumulation,
increased
mitosis,
and/
or
focal
liver
cell
necrosis
in
males
at
=
4.5
mg/
kg/
day
and
females
at
=
5
mg/
kg/
day;
and
centrilobular
hepatocytic
hypertrophy,
proliferation
of
smooth
endoplasmic
reticulum,
increased
microsomal
activity,
and/
or
increased
glycogen
content
in
males
at
22.5
mg/
kg/
day
and
females
at
25
mg/
kg/
day.
Other
systemic
effects
included
increased
absolute
and/
or
kidney
weight
in
females
at
=
2.0
mg/
kg/
day
and
males
at
=
4.5
mg/
kg/
day;
renal
medulla
calcinosis
in
males
at
22.5
mg/
kg/
day;
and
clinical
signs
(
ataxia
progressing
to
inactivity
and
coma),
hematologic
and
splenic
changes
indicative
of
anemia
(
decreased
red
blood
cells
and
hemoglobin,
increased
extramedullar
hematopoiesis),
and
reduced
body
weight
in
males
at
22.5
mg/
kg/
day
and
females
at
25
mg/
kg/
day.
Due
to
the
dose­
related
nature
and
progression
in
severity
of
the
hepatic
effects,
and
the
mild,
reversible
nature
of
the
changes
at
the
lowest
dose
level,
0.18
mg/
kg/
day
is
considered
to
be
a
minimal
LOAEL
based
on
hyalinization
of
centrilobular
cells.
The
liver
is
an
established
target
of
ß­
HCH
in
other
subchronic
and
chronic
studies
in
rats
and
mice
(
Fitzhugh
et
al.
1950;
Ikegami
et
al.
1991a,
1991b;
Ito
et
al.
1973;
Schoter
et
al.
1987).

RIVM
­
RIVM
derived
chronic
oral
TDI
of
RfD
of
0.00002
for
$­
HCH
based
on
a
NOAEL
of
0.02
mg/
kg­
day
for
observations
of
infertility
in
two
semi­
chronic
oral
studies
on
reproduction
in
rats.
RIVM
applied
an
UF
of
1000
(
10
each
to
account
for
inter­
and
intraspecies
variability
and
a
poor
database)
(
Slooff
and
Matthijsen,
1988).

Cancer
ATSDR
and
U.
S.
EPA
have
evaluated
the
carcinogenicity
data
for
$­
HCH.
IRIS
classifies
this
chemical
as
C­
possible
human
carcinogen
and
established
an
estimated
oral
slope
or
cancer
potency
factor
of
1.8
(
mg/
kg/
day)­
1.
The
cancer
weight­
of­
evidence
classification
is
based
on
all
routes
of
exposure.

The
IRIS
cancer
classification
is
based
on
increases
in
benign
liver
tumors
in
CF1
mice
fed
$­
HCH.
Positive
or
marginally
positive
tumorigenic
responses,
characterized
as
benign
hepatomas
or
hepatocellular
carcinomas,
have
been
observed
in
two
strains
of
mice.
The
studies
are
limited
in
that
small
numbers
of
animals
were
used,
no
doseresponse
data
are
available,
not
all
of
the
animals
were
examined
histologically,
or
the
duration
of
exposure
was
less
than
lifetime.
Thorpe
and
Walker
(
1973)
fed
30
each
male
and
female
CF1
mice
dietary
$­
HCH
at
200
ppm
for
110
weeks.
This
resulted
in
12%
mortality
of
males
and
25%
of
females
during
the
first
3
months.
A
significantly
increased
incidence
of
liver
tumors
was
observed
in
treated
males
and
females.
No
statistically
significant
evidence
of
increased
tumor
incidence
as
a
consequence
of
$­
HCH
feeding
was
seen
in
several
small
(
5­
20/
group)
studies
with
male
and
female
dd
mice
fed
0­
600
ppm
for
24­
32
weeks
(
Nagasaki
et
al.,
1972;
Hanada
et
al.
1973;
Ito
et
al.,
1973)
or
in
male
Wistar
rats
(
Ito
et
al.,
1975;
Fitzhugh
et
al.,
1950)
fed
0­
1000
ppm
for
>
72
weeks.
Goto
et
al.
(
1972)
maintained
male
ICR­
JCL
mice
on
a
diet
containing
600
45
ppm
$­
HCH
for
26
weeks.
Relative
liver
weight
was
increased
in
the
treated
animals,
and
there
was
histologic
evidence
of
benign
neoplasms
at
an
unspecified
incidence.
Table
9
summarizes
the
established
toxicological
endpoints
for
beta­
HCH.

Table
9.
Beta­
HCH
Summary
of
Doses
and
Toxicological
Endpoints
Risk
Assessment
Exposure
Route
Dose/
Endpoint
Endpoint
Source
Acute
Oral
NOAEL
=
4.5
mg/
kg­
day
UF
=
100
RfD
=
0.05
mg/
kg­
day*
LOAEL
=
22.5
mg/
kgday
based
on
clinical
signs
of
ataxia
in
rats
Van
Velsen
et
al.
1986
(
ATSDR,
2005)

Chronic
Oral
LOAEL
=
0.18
mg/
kg­
day
UF
=
3000
RfD
=
0.00006
mg/
kg­
day*
LOAEL
=
0.18
mg/
kgday
based
on
liver
effects
in
rats
Van
Velsen
et
al.
1986
(
ATSDR,
2005)

Chronic
Oral
NOAEL=
of
0.02
mg/
kg/
day
UF
=
1000
RfD
=
0.00002
mg/
kg/
day
LOAEL
=
NP
observations
of
infertility
in
a
reproduction
toxicity
assay
in
rats
Slooff
and
Matthijsen,
1988
(
RIVM
2001)

Cancer
(
all
routes)
Classification
­
C;
possible
human
carcinogen
(
EPA
IRIS
1993)
Q1*
=
1.8
(
mg/
kg/
day)­
1
*
Basis
­­
Increases
in
benign
liver
tumors
in
CF1
mice
fed
$­
HCH
(
Thorpe
and
Walker
1973).
*
Endpoint
used
in
dietary
risk
assessment
5.
Dietary
Risk
Assessment
The
Agency
performed
dietary
risk
and
exposure
assessments
to
estimate
risk
to
the
Indigenous
People
from
exposure
to
the
"­
and
$­
HCH
isomers
using
data
on
subsistence
food
harvest
amounts,
total
HCH
residues
in
traditional
foods,
and
estimated
ranges
of
"­
and
$­
HCH
as
a
proportion
of
total
HCH
residues.
Only
Chronic
(
noncancer
and
cancer)
dietary
exposures
were
evaluated
for
"­
HCH
based
on
available
toxicity
data.
Both
acute
and
chronic
(
non­
cancer
and
cancer)
dietary
exposures
were
evaluated
for
$­
HCH
based
on
available
toxicity
data.

a.
Dietary
Profile
The
Agency
used
the
subsistence
food
harvest
amounts
of
nearly
180
communities
from
the
Community
Profile
Database
Version
3.11
dated
3/
27/
01
from
the
Alaska
Department
of
Fish
and
Game
Division
of
Subsistence
as
subsistence
food
intake
rates.
This
database
includes
the
harvest
of
subsistence
foods
in
Alaskan
communities
from
the
years
1990­
2001.
From
personal
communication
with
Mr.
Roland
Shanks
from
the
Alaska
Inter­
Tribal
Council,
it
was
determined
that
usually
walrus,
seal,
and
whale
are
the
marine
mammals
harvested
in
significant
amounts.
Therefore,
the
Agency
used
data
from
the
communities
with
the
highest
representative
seal,
walrus
whale
harvests
to
develop
conservative
estimates
of
subsistence
dietary
intake
for
the
dietary
risk
assessment.
Harvest
data
on
polar
bear,
other
marine
mammal,
fish,
other
land
mammals,
46
and
birds
from
the
corresponding
Alaskan
community
were
also
included
in
EPA's
estimate
of
subsistence
dietary
consumption.
EPA
used
per
capita
harvest
amounts
as
the
human
intake
amount.
This
results
in
a
conservative
estimate
since
some
of
the
harvest
is
likely
used
for
non­
human
food
purposes
and
some
disposed
of
as
waste.
Estimated
subsistence
intake
for
the
three
communities
with
the
highest
walrus,
seal,
and
whale
harvest
are
provided
in
Table
11.

EPA
coupled
this
dietary
intake
data
with
organochlorine
residue
data
obtained
from
Dr.
Laurie
Chan
of
McGill
University
in
Canada
via
personal
communication.
The
data
gave
analytical
results
of
the
samples
for
total
HCH
in
numerous
traditional
foods
as
shown
in
Table
11.
Information
on
the
individual
isomers
was
not
provided.
The
Agency
used
subsistence
food
harvest
amounts
and
total
HCH
residues
in
traditional
foods
for
Community
1
(
the
community
with
the
highest
total
HCH
exposure
of
the
three
communities)
to
estimate
subsistence
exposure.
Based
on
these
data,
adults
are
assumed
to
consume
up
to
2.4
pounds
and
children
up
to
1.3
pounds
of
subsistence
meat
per
day.
The
adult
intake
amounts
were
adjusted
by
a
factor
of
0.53
to
obtain
the
intake
amount
for
children;
this
factor
was
derived
from
published
data
(
Heller,
1966)
(
Table
10).
Exposure
to
total
HCH
for
Community
1
is
estimated
at
282,065
ng/
day
(
Table
11).

Table
10.
Subsistence
Meat
Consumption
by
Community
Community
Population
Pt.
Hope
Notak
Shungnak
Adult
Males
438
grams
429
grams
573
grams
Children
7­
12
years
old
230
grams
230
grams
303
grams
Child's
%
of
Adults
Subsistence
Meat
Consumption
52.5
53.6
52.9
Table
11.
Community
Harvest
of
Traditional
Foods
and
Total
HCH
Residues
Traditional
Food
Total
HCH
Residues
(
ng/
g)
Community
1
Harvest
(
grams/
person/
day)
Community
2
Harvest
(
grams/
person/
day)
Community
3
Harvest
(
grams/
person/
day)

Polar
Bear
10
9
26
16
Seal
215
39
500
46
Whale
391
697
­­­­
271
Walrus
20
­­­­
22
315
Caribou
1
123
103
221
Moose
9
­­­­
81
­­­­

Muskox
2
13
­­­­
­­­­

Dall
Sheep
4
20
­­­­
­­­­

Salmon
26
28
116
­­­­

Arctic
Char
6
­­­­
10
­­­­

Lake
Trout
3
100
­­­­
­­­­
47
Table
11.
Community
Harvest
of
Traditional
Foods
and
Total
HCH
Residues
Traditional
Food
Total
HCH
Residues
(
ng/
g)
Community
1
Harvest
(
grams/
person/
day)
Community
2
Harvest
(
grams/
person/
day)
Community
3
Harvest
(
grams/
person/
day)

Arctic
Grayling
3
6
­­­­
6
Whitefish
20
­­­­
14
19
Cod
­­­­
18
­­­­

Smelt
­­­­
10
17
Herring
Residue
from
ooligan
flesh
used
­
348
ng/
g
­­­­
22
­­­­

Cisco
1
39
8
19
Goose
1
14
15
14
Duck
7
­­­­
12
­­­­

Berries
10
­­­­
15
­­­­

Total
HCH
Exposure
282,065
ng/
day
130,045
ng/
day
128,879
ng/
day
The
Agency
adjusted
the
total
HCH
exposure
based
on
the
proportion
of
"­
and
$­
HCH
likely
to
be
present
in
total
HCH.
"­
HCH
represents
between
31
and
97%
of
total
HCH
residues.
$­
HCH
represents
between
20
and
69%
of
total
HCH
residues.
These
factors
were
derived
from
published
articles
in
which
"­
HCH,
$­
HCH,
and
total
HCH
were
measured
(
Kucklick,
et
al.
2002;
Lee,
et
al.
1996;
Metcalfe,
et
al.
1999;
Hoekstra,
et
al.
2005).
Assuming
body
weights
of
70
kg
for
males,
60
kg
for
females,
and
10­
30
kg
for
children,
EPA
estimated
"
and
$
exposures
in
the
range
of
0.00057
­
0.0039
and
0.00037
­
0.0028
mg/
kg
bw/
day,
respectively,
for
adult
male
indigenous
people;
0.00067
­
0.0046
and
0.00043
­
0.0032
mg/
kg
bw/
day
for
adult
females;
0.0021
­
0.051
and
0.0014
­
0.010
mg/
kg
bw/
day
for
children
1­
6;
and
0.00073­
0.0050
and
0.00048­
0.0036
mg/
kg
bw/
day
for
children
7­
12.

b.
Dietary
Exposure
and
Risk
Dietary
risk
assessment
incorporates
both
exposure
and
toxicity
of
a
given
pesticide.
For
acute
and
chronic
assessments,
the
risk
is
expressed
as
a
percentage
of
a
maximum
acceptable
dose
or
RfD
(
i.
e.,
the
dose
which
EPA
has
concluded
will
result
in
no
unreasonable
adverse
health
effects).
For
acute
and
non­
cancer
chronic
exposures,
EPA
is
concerned
when
estimated
dietary
risk
exceeds
100%
of
the
RfD.
The
Agency
is
generally
concerned
when
estimated
cancer
risk
exceeds
one
in
one
million
(
i.
e.,
the
risk
exceeds
1
x
10­
6).
EPA's
dietary
risk
assessment
indicates
that
the
chronic
dietary
exposure
estimates
for
$­
HCH
are
above
EPA's
level
of
concern
for
both
low
and
highend
dietary
intake
estimates;
chronic
dietary
exposure
estimates
for
"­
HCH
are
above
EPA's
level
of
concern
for
high­
end
exposure
estimates.
The
cancer
dietary
risk
estimates
for
both
"­
and
$­
HCH
are
above
the
Agency's
level
of
concern
for
both
low
and
high­
end
dietary
intake
estimates.
The
acute
dietary
exposure
estimates
for
$­
HCH
48
are
below
EPA's
level
of
concern.
The
results
of
the
dietary
exposure
analyses
are
reported
in
the
Tables
12,
13,
and
14.

Table
12.
Estimated
Total
Dietary
Intake
of
beta­
HCH
and
Acute
Risk.

Population
Subgroup
Body
Weight
(
kg)
Estimated
beta­
HCH
Exposure
(
mg/
kg/
day)
%
aRfD*

Adult
male
70
0.00037
­
0.0028
<
1­
4
Adult
female
60
0.00043
­
0.0032
<
1­
8
Children
(
1­
6
yrs)
10
0.0014
­
0.010
<
1­
20
Children
(
7­
12
yrs.)
29
0.00048­
0.0036
<
1­
8
*
aRfD
=
0.05
mg/
kg/
day
Table
13.
Estimated
Total
Dietary
Intake
of
alpha­
HCH
and
Chronic/
Cancer
Risk
Population
Subgroup
Body
Weight
(
kg)
Estimated
alpha­
HCH
Exposure
(
mg/
kg/
day)
%
cRfD*
Cancer
Risk**

Adult
male
70
0.00057
­
0.0039
57­
390
3.6x10­
3
to
2.5x10­
2
Adult
female
60
0.00067
­
0.0046
67­
460
4.2x10­
3
to
2.9x10­
2
Children
(
1­
6
yrs)
10
0.0021
­
0.051
210­
5100
NA
Children
(
7­
12
yrs.)
29
0.00073­
0.0050
73­
500
NA
*
cRfD
=
0.001
mg/
kg/
day
**
Q1*
=
6.3
mg/
kg/
day­
1
Table
14.
Estimated
Total
Dietary
Intake
of
beta­
HCH
and
Chronic/
Cancer
Risk
Population
Subgroup
Body
Weight
(
kg)
Estimated
beta­
HCH
Exposure
(
mg/
kg/
day)
%
cRfD*
Cancer
Risk**

Adult
male
70
0.00037
­
0.0028
620­
4700
6.7x10­
4
to
5.0x10­
3
Adult
female
60
0.00043
­
0.0032
720­
5300
7.7x10­
4
to
5.8x10­
3
Children
(
1­
6
yrs)
10
0.0014
­
0.010
2300­
17000
NA
Children
(
7­
12
yrs.)
29
0.00048­
0.0036
800­
6000
NA
*
cRfD
=
0.00006
mg/
kg/
day
**
Q1*
=
1.8
mg/
kg/
day­
1
c.
Risk
Characterization
The
Agency
believes
this
is
a
conservative
dietary
risk
assessment
likely
to
overestimate
dietary
exposure
and
risk
since
high­
end
assumptions
were
used
for
several
critical
or
risk
driving
parameters.
For
example,
dietary
intake
was
assumed
to
be
equal
to
harvest.
This
does
not
take
into
account
portions
of
the
harvest
which
were
discarded
or
used
for
non­
dietary
purposes.
Also,
the
maximum
detected
HCH
residue
concentration
was
used
in
the
calculation
of
the
HCH
exposure
in
the
respective
49
subsistence
food
item
(
e.
g.,
the
beluga
whale
blubber
HCH
concentration
was
assumed
for
the
entire
harvest
(
and
consumption)
amount).
This
does
not
reflect
the
fact
that
other
tissues
consumed
may
have
had
much
lower
residue
amounts.
In
addition,
it
is
assumed
that
a
subsistence
diet
is
consumed
year
round.
In
fact
high
subsistence
intake
is
more
likely
to
be
seasonal.
It
is
also
important
to
reiterate
the
likelihood
that
intake
rates
among
tribal
communities
in
Alaska
may
vary
greatly
depending
on
the
availability
of
fish
and
game
meat.
Nevertheless,
EPA
believes
that
use
of
less
conservative
assumptions
(
e.
g.,
dietary
intake
equal
to
half
the
harvest,
substantially
lower
subsistence
meat
consumed
per
day)
would
still
result
in
indigenous
dietary
risks
well
above
the
Agency's
level
of
concern.
It
is
also
important
to
note
that
both
the
cancer
and
noncancer
endpoints
selected
for
exposure
to
"­
and
$­
HCH
are
based
on
liver
effects.
Therefore,
exposures
to
"­
and
$­
HCH
may
be
additive.
Moreover,
the
chronic
noncancer
endpoint
for
exposure
to
lindane,
the
gamma
((­)
isomer,
is
also
based
on
liver
effects.
The
Agency
is
inviting
comment.
50
V.
Additional
Concerns
and
Information
Request
Additional
concerns
related
to
lindane
and
the
HCH
isomers
have
been
raised
in
public
comments
on
the
lindane
RED
and
risk
assessments
(
Docket
Numbers
OPP­
2002­
0202
and
OPP
34239)
and
comments
on
the
draft
NARAP.
The
Agency
would
like
to
obtain
additional
information
from
the
public
specific
to
the
topics
listed
below
as
it
makes
its
final
determination
on
lindane.

A.
Infants'
exposure
to
lindane
and
the
HCH
isomers
in
breast
milk
(
general
population
and
subsistence
populations)

The
Agency
would
like
to
receive
information
on
lindane
levels,
as
well
as
levels
of
other
HCH
isomers,
in
breast
milk.

B.
Cancer
classification
In
comments,
the
public
has
expressed
disagreement
with
the
Agency's
cancer
classification
for
lindane
as
presented
in
the
lindane
RED.
The
Agency
would
like
to
obtain
additional
information
that
may
be
available
on
the
carcinogenicity
of
lindane.

C.
10x
FQPA
Safety
Factor
In
several
comments,
the
public
has
disagreed
with
EPA's
rationale
for
reducing
the
FQPA
ten­
fold
safety
factor
to
3x
for
lindane.
The
Agency
would
like
to
receive
additional
information
on
the
rationale
presented
by
commenters
for
retaining
the
10x
FQPA
safety
factor
on
lindane.

D.
Cultural
Practices
and
Potential
Impacts
to
Subsistence
Populations
The
Agency
received
several
comments
from
the
public
on
the
dietary
exposure
assessment
conducted
in
support
of
the
RED
for
the
indigenous
people
of
Alaska,
particularly
the
dietary
intake
assumptions.
EPA
seeks
to
achieve
environmental
justice,
the
fair
treatment
and
meaningful
involvement
of
all
people,
regardless
of
race,
color,
national
origin,
or
income,
in
the
development,
implementation,
and
enforcement
of
environmental
laws,
regulations,
and
policies.
To
help
address
potential
environmental
justice
issues,
the
Agency
seeks
information
on
any
groups
or
segments
of
the
population
who,
as
a
result
of
their
location,
cultural
practices,
or
other
factors,
may
have
atypical,
unusually
high
exposure
to
lindane
and
the
other
HCH
isomers,
compared
to
the
general
populations.
As
such,
the
Agency
would
like
to
receive
additional
information
on
actual
dietary
intake
and
other
practices
taking
place
in
Alaskan
subsistence
cultures
that
may
impact
the
assumptions
used
in
this
assessment
of
lindane
and
the
other
HCH
isomers.

E.
Liver
Effects
Both
the
cancer
and
non­
cancer
endpoints
selected
for
exposure
to
"­
and
$­
HCH
are
based
on
liver
effects.
Therefore,
exposures
to
"­
and
$­
HCH
may
be
additive.
51
Moreover,
the
chronic
non­
cancer
endpoint
for
exposure
to
lindane,
the
gamma
((­)
isomer,
is
also
based
on
liver
effects.
The
Agency
is
inviting
comment.
52
VI.
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Appendix
A
 
Summary
List
of
International
Lindane
Registration
Status
by
Country
Banned
Argentina
Armenia
Banladesh
Barbados
Belgium
Bulgaria
Burundi
Costa
Rica
Croatia
Cyrus
Czech
Republic
Denmark
Dominican
Republic
Ecuador
Egypt
El
Salvador
Finland
Gambia
Georgia
Guatemala
Honduras
Hong
Kong
Hungary
Jamaica
Japan
Kazakhstan
Korea,
Dem.
Rep
Korea,
Rep
Latvia
Liechtenstein
Lithuania
Mozambique
Netherlands
New
Zealand
Nicaragua
Norway
Paraguay
Peru
Poland
Information
for
this
table
was
taken
from
the
document
"
International
Registration
Status
of
Lindane
by
Country"
found
at:
www.
cec.
org
Russia
(?)
Singapore
Slovakia
South
Africa
St
Lucia
Sweden
Taiwan
Thailand
Tonga
Turkey
Uruguay
Vietnam
Yemen
Restricted/
Severely
Restricted
Algeria
Australia
Austria
Belize
Brazil
Canada
China
Columbia
Cuba
European
Community
Fiji
France
Germany
Iceland
Ireland
Israel
Italy
Madagascar
Moldova
Morocco
Nigeria
Philippines
Samoa
Senegal
Spain
Sri
Lanka
Sudan
Switzerland
Trinidad/
Tobago
United
Kingdom
United
Status
of
America
Venezuela
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Not
registered
Estonia
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Bissau
Indonesia
Monaco
Mongolia
Niger
Rwanda
Slovenia
Uganda
Vanuatu
Registered
Bolivia
Burkina
Faso
Cameroon
Cape
Verde
Chad
India
Kenya
Malaysia
Mali
Mauritania
Mexico
Papua
New
Guinea
Portugal
Syria
Tanzania
Togo
Zimbabwe
