Document ID: EPA-HQ-OW-2002-0021-0132
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2003-05-15T04:00Z

Health
Effects
Support
Document
for
Aldrin/
Dieldrin
Printed
on
Recycled
Paper
Health
Effects
Support
Document
for
Aldrin/
Dieldrin
Prepared
for:

U.
S.
Environmental
Protection
Agency
Office
of
Water
(
4304T)
Health
and
Ecological
Criteria
Division
Washington,
DC
20460
www.
epa.
gov/
safewater/
ccl/
pdf/
aldrindieldrin.
pdf
EPA
822­
R­
03­
001
February
2003
iii
Aldrin/
Dieldrin
 
February
2003
FOREWORD
The
Safe
Drinking
Water
Act
(
SDWA),
as
amended
in
1996,
requires
the
Administrator
of
the
Environmental
Protection
Agency
to
establish
a
list
of
contaminants
to
aid
the
agency
in
regulatory
priority
setting
for
the
drinking
water
program.
In
addition,
SDWA
requires
EPA
to
make
regulatory
determinations
for
no
fewer
than
five
contaminants
by
August
2001.
The
criteria
used
to
determine
whether
or
not
to
regulate
a
chemical
on
the
CCL
are
as
follows:

The
contaminant
may
have
an
adverse
effect
on
the
health
of
persons.

The
contaminant
is
known
to
occur
or
there
is
a
substantial
likelihood
that
the
contaminant
will
occur
in
public
water
systems
with
a
frequency
and
at
levels
of
public
health
concern.

In
the
sole
judgment
of
the
administrator,
regulation
of
such
contaminant
presents
a
meaningful
opportunity
for
health
risk
reduction
for
persons
served
by
public
water
systems.

The
Agency's
findings
for
all
three
statutory
criteria
are
used
in
order
to
make
a
determination
to
regulate
a
contaminant.
The
Agency
may
determine
that
there
is
no
need
for
a
regulation
when
a
contaminant
fails
to
meet
one
of
the
statutory
criteria.
A
decision
not
to
regulate
is
considered
a
final
agency
action
and
is
subject
to
judicial
review.

This
document
provides
the
health
effects
basis
for
the
preliminary
regulatory
determination
for
aldrin
and
dieldrin.
In
arriving
at
the
preliminary
regulatory
determination
for
these
two
contaminants,
data
on
toxicokinetics,
human
exposure,
acute
and
chronic
toxicity
to
animals
and
humans,
epidemiology,
and
mechanisms
of
toxicity
were
evaluated.
In
order
to
avoid
wasteful
duplication
of
effort,
information
from
the
following
risk
assessments
by
the
EPA
and
other
government
agencies
were
used
in
development
of
this
document.

ATSDR.
2000.
Agency
for
Toxic
Substances
and
Disease
Registry.
Draft
Toxicological
Profile
for
Aldrin/
Dieldrin:
Update.
Atlanta,
GA:
U.
S.
Department
of
Health
and
Human
Services.

ATSDR.
1993.
Agency
for
Toxic
Substances
and
Disease
Registry.
Toxicological
Profile
for
Aldrin/
Dieldrin.
Atlanta,
GA:
USDepartment
of
Health
and
Human
Services.

USEPA.
1992.
US
Environmental
Protection
Agency.
Aldrin
Drinking
Water
Health
Advisory.
Office
of
Water.

USEPA.
1988.
US
Environmental
Protection
Agency.
Dieldrin
Drinking
Water
Health
Advisory.
Office
of
Water.
iv
Aldrin/
Dieldrin
 
February
2003
USEPA.
1987a.
US
Environmental
Protection
Agency.
Integrated
Risk
Information
System
(
IRIS):
Dieldrin.
Cincinnati,
OH.

USEPA.
1987b.
US
Environmental
Protection
Agency.
Carcinogenicity
assessment
of
Dieldrin
and
Aldrin.
(
CAG).

USEPA.
1986.
US
Environmental
Protection
Agency.
Integrated
Risk
Information
System
(
IRIS):
Aldrin.
Cincinnati,
OH.

IARC.
1987.
International
Agency
for
Research
on
Cancer.
Evaluation
of
the
carcinogenic
risk
of
chemicals
to
humans.
Overall
evaluations
of
carcinogenicity.
Suppl.
7:
88­
89.

IARC.
1982.
International
Agency
for
Research
on
Cancer.
IARC
monographs
on
the
evaluation
of
the
carcinogenic
risk
of
chemicals
to
humans.
Chemicals,
industry
process
and
industries
associated
with
cancer
in
humans.
IARC
Monographs.
Vols.
1­
29,
Supplement
4.
Geneva:
World
Health
Organization.

IARC.
1974a.
International
Agency
for
Research
on
Cancer.
Evaluation
of
the
carcinogenic
risk
of
chemicals
to
humans.
Aldrin.
Lyon,
France:
IARC
Monograph
5:
25­
38.

IARC.
1974b.
International
Agency
for
Research
on
Cancer.
Evaluation
of
the
carcinogenic
risk
of
chemicals
to
humans.
Dieldrin.
Lyon,
France:
IARC
Monograph
5:
125­
156.

In
cases
where
the
information
in
this
document
originates
from
one
of
the
references
above,
a
citation
to
the
source
document
is
provided
with
the
bibliographic
information
in
the
reference
section.
Primary
references
were
used
for
all
key
studies.
Data
from
the
published
risk
assessments
were
supplemented
with
information
from
literature
searches
conducted
in
2000.
Specific
emphasis
is
placed
on
dose­
response
information
and
exposure
estimates
in
making
the
regulatory
determination
for
aldrin
and
dieldrin.
Dose­
reponse
conclusions
for
noncancer
effects
are
reflected
in
the
Reference
Dose
(
RfD).

Generally,
a
RfD
is
provided
as
the
assessment
of
long­
term
toxic
effects
other
than
carcinogenicity.
RfD
determination
assumes
that
thresholds
exist
for
certain
toxic
effects,
such
as
cellular
necrosis.
It
is
expressed
in
terms
of
milligrams
per
kilogram
per
day
(
mg/
kg­
day).
In
general,
the
RfD
is
an
estimate
(
with
uncertainty
spanning
perhaps
an
order
of
magnitude)
of
a
daily
exposure
to
the
human
population
(
including
sensitive
subgroups)
that
is
likely
to
be
without
an
appreciable
risk
of
deleterious
effects
during
a
lifetime.

The
carcinogenicity
assessment
for
aldrin
and
dieldrin
includes
a
formal
hazard
identification.
Hazard
identification
is
a
weight­
of­
evidence
judgement
of
the
likelihood
that
the
agent
is
a
human
carcinogen
via
the
oral
route
and
the
conditions
under
which
the
carcinogenic
effects
may
be
expressed.
v
Aldrin/
Dieldrin
 
February
2003
Guidelines
that
were
used
in
the
development
of
this
assessment
may
include
the
following:
the
Guidelines
for
Carcinogen
Risk
Assessment
(
USEPA,
1986a),
Guidelines
for
the
Health
Risk
Assessment
of
Chemical
Mixtures
(
USEPA,
1986b),
Guidelines
for
Mutagenicity
Risk
Assessment
(
USEPA,
1986c),
Guidelines
for
Developmental
Toxicity
Risk
Assessment
(
USEPA,
1991),
Proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
1996a),
Guidelines
for
Reproductive
Toxicity
Risk
Assessment
(
USEPA,
1996b),
and
Guidelines
for
Neurotoxicity
Risk
Assessment
(
USEPA,
1998a);
Recommendations
for
and
Documentation
of
Biological
Values
for
Use
in
Risk
Assessment
(
USEPA,
1988);
and
Health
Effects
Testing
Guidelines
(
OPPTS
series
870,
1996
drafts;
USEPA
40
CFR
Part
798,
1997;
Peer
Review
and
Peer
Involvement
at
the
U.
S.
Environmental
Protection
Agency
(
USEPA,
1994c);
Use
of
the
Benchmark
Dose
Approach
in
Health
Risk
Assessment
(
USEPA,
1995b);
Science
Policy
Council
Handbook:
Peer
Review
(
USEPA,
1998b,
2000a);
Memorandum
from
EPA
Administrator,
Carol
Browner,
dated
March
21,
1995,
Policy
for
Risk
Characterization;
Science
Policy
Council
Handbook:
Risk
Characterization
(
USEPA,
2000b).

The
section
on
aldrin
and
dieldrin
occurrence
and
exposure
through
potable
water
in
this
document
was
developed
by
the
Office
of
Ground
Water
and
Drinking
Water.
It
is
based
primarily
on
unregulated
contaminant
monitoring
(
UCM)
data
collected
under
SDWA.
The
UCM
data
are
supplemented
with
ambient
water
data,
as
well
as
information
on
production,
use,
and
discharge.
vi
Aldrin/
Dieldrin
 
February
2003
ACKNOWLEDGMENTS
This
document
was
prepared
under
U.
S.
EPA
Contract
No.
68­
C­
01­
002,
Work
Assignment
No.
B­
11,
with
Sciences
International,
Alexandria,
VA.
The
Lead
Scientist
is
Amal
M.
Mahfouz,
Ph.
D.,
Health
and
Ecological
Criteria
Division,
Office
of
Science
and
Technology,
Office
of
Water.
vii
Aldrin/
Dieldrin
 
February
2003
TABLE
OF
CONTENTS
FOREWORD
.
.
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iii
ACKNOWLEDGMENT
.
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vi
LIST
OF
TABLES
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x
LIST
OF
FIGURES
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xi
1.0
EXECUTIVE
SUMMARY
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1­
1
2.0
IDENTITY:
PHYSICAL
AND
CHEMICAL
PROPERTIES
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2­
1
3.0
USES
AND
ENVIRONMENTAL
FATE
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3­
1
3.1
Uses
and
Manufacture
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3­
1
3.2
Environmental
Release
and
Fate
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.
3­
2
4.0
EXPOSURE
FROM
DRINKING
WATER
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.
4­
1
4.1
ALDRIN
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.
4­
1
4.1.1
Ambient
Occurrence
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4­
1
4.1.2
Drinking
Water
Occurrence
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4­
3
4.1.3
Conclusion
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4­
13
4.2
DIELDRIN
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4­
19
4.2.1
Ambient
Occurrence
.
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4­
19
4.2.2
Drinking
Water
Occurrence
.
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4­
22
4.2.3
Conclusion
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4­
35
5.0
EXPOSURE
FROM
ENVIRONMENTAL
MEDIA
OTHER
THAN
WATER
.
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5­
1
5.1
Exposure
from
Food
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5­
1
5.1.1
Exposures
of
the
General
Population
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5­
1
5.1.2
Exposures
of
Subpopulations
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5­
7
5.2
Exposure
from
Air
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5­
8
5.2.1
Exposures
of
the
General
Population
.
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5­
8
5.2.2
Exposures
of
Subpopulations
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5­
11
5.3
Exposure
from
Soil
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5­
11
5.3.1
Exposures
of
the
General
Population
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5­
11
5.3.2
Exposures
of
Subpopulations
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5­
12
5.4
Other
Residential
Exposures
(
Not
Drinking
Water
Related)
.
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.
.
5­
13
5.5
Summary
of
Exposure
to
Aldrin/
Dieldrin
in
Media
Other
Than
Water
.
.
.
.
.
5­
15
6.0
TOXICOKINETICS
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6­
1
6.1
Absorption
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6­
1
viii
Aldrin/
Dieldrin
 
February
2003
6.2
Distribution
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6­
2
6.3
Metabolism
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6­
11
6.4
Excretion
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6­
17
7.0
HAZARD
IDENTIFICATION
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7­
1
7.1
Human
Effects
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7­
1
7.1.1
Short­
Term
Studies
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7­
1
7.1.2
Long­
Term
and
Epidemiological
Studies
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7­
2
7.2
Animal
Studies
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7­
5
7.2.1
Acute
Toxicity
(
Oral,
Inhalation,
Dermal)
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7­
5
7.2.2
Short­
Term
Studies
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7­
6
7.2.3
Subchronic
Studies
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7­
7
7.2.4
Neurotoxicity
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7­
9
7.2.5
Developmental/
Reproductive
Toxicity
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7­
12
7.2.6
Chronic
Toxicity
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7­
16
7.2.7
Carcinogenicity
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7­
20
7.3
Other
Key
Data
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7­
25
7.3.1
Mutagenicity/
Genotoxicity
Effects
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7­
25
7.3.2
Immunotoxicity
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7­
27
7.3.3
Hormonal
Disruption
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7­
28
7.3.4
Physiological
or
Mechanistic
Studies
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7­
28
7.3.5
Structure­
Activity
Relationship
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7­
32
7.4
Hazard
Characterization
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7­
33
7.4.1
Synthesis
and
Evaluation
of
Noncancer
Effects
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7­
33
7.4.2
Synthesis
and
Evaluation
of
Carcinogenic
Effects
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7­
34
7.4.3
Mode
of
Action
and
Implications
in
Cancer
Assessment
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7­
36
7.4.4
Weight
of
Evidence
Evaluation
for
Carcinogenicity
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7­
38
7.4.5
Sensitive
Populations
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7­
39
8.0
DOSE­
RESPONSE
ASSESSMENTS
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8­
1
8.1
Dose­
Response
for
Non­
Cancer
Effects
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8­
1
8.1.1
Reference
Dose
Determination
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8­
1
8.1.2
Reference
Concentration
(
RfC)
Determination
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8­
3
8.2
Dose­
Response
for
Cancer
Effects
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8­
3
8.2.1
Choice
of
Study/
Data
With
Rationale
and
Justification
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8­
3
9.0
REGULATORY
DETERMINATION
AND
CHARACTERIZATION
OF
RISK
FROM
DRINKING
WATER
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9­
1
9.1
Regulatory
Determination
for
Chemicals
on
the
CCL
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9­
1
9.1.1
Criteria
for
Regulatory
Determination
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9­
1
9.1.2
National
Drinking
Water
Advisory
Council
Recommendations
.
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9­
2
9.2
Health
Effects
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9­
2
9.2.1
Health
Criterion
Conclusions
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9­
3
9.2.2
Hazard
Characterization
and
Mode
of
Action
Implications
.
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9­
3
ix
Aldrin/
Dieldrin
 
February
2003
9.2.3
Dose­
Response
Characterization
and
Implications
in
Risk
Assessment
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9­
6
9.3
Occurrence
in
Public
Water
Systems
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9­
12
9.3.1
Occurrence
Criterion
Conclusions
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9­
12
9.3.2
Monitoring
Data
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9­
13
9.3.3
Use
and
Fate
Data
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9­
16
9.4
Risk
Reduction
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9­
17
9.4.1
Risk
Criterion
Conclusions
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9­
17
9.4.2
Exposed
Population
Estimates
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9­
17
9.4.3
Relative
Source
Contribution
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9­
19
9.4.4
Sensitive
Populations
.
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9­
20
9.5
Regulatory
Determination
Summary
.
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9­
20
10.0
REFERENCES
.
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10­
1
Abbreviations
and
Acronyms
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.
A1
APPENDIX
A:
Round
2
Aldrin
Occurrence
.
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.
B1
APPENDIX
B:
Round
2
Dieldrin
Occurrence
.
.
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.
B2
x
Aldrin/
Dieldrin
 
February
2003
LIST
OF
TABLES
Table
2­
1.
Selected
Chemical­
Physical
Properties
of
Aldrin
and
Dieldrin
.
.
.
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.
.
2­
3
Table
3­
1.
Aldrin
Mobility
in
Soils
Used
to
Grow
Corn
.
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.
3­
4
Table
4­
1.
Aldrin
Detections
in
Stream
Bed
Sediments
.
.
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.
.
4­
3
Table
4­
2.
Summary
Occurrence
Statistics
for
Aldrin
.
.
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.
4­
12
Table
4­
3.
Dieldrin
Detections
and
Concentrations
in
Streams
and
Ground
Water
.
.
.
.
.
4­
21
Table
4­
4.
Dieldrin
Detections
and
Concentrations
in
Sediments,
Whole
Fish,
and
Bivalves
(
All
Sites)
.
.
.
.
.
.
.
.
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.
4­
22
Table
4­
5.
Summary
Occurrence
Statistics
for
Dieldrin
.
.
.
.
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.
4­
32
Table
5­
1.
Aldrin
and
Dieldrin
in
Domestic
Food
Items
1981
to
1992
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
5­
2
Table
5­
2.
Aldrin
Concentrations
in
San
Francisco
Bay
Area
Fish
in
1994
.
.
.
.
.
.
.
.
.
.
.
.
5­
6
Table
5­
3.
Summary
of
General
Population
Exposures
to
Aldrin
in
Media
Other
than
Water
.
.
.
.
.
.
.
.
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.
.
.
5­
15
Table
5­
4.
Summary
of
General
Population
Exposure
to
Dieldrin
in
Media
Other
than
Water
.
.
.
.
.
.
.
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.
.
5­
16
Table
5­
5.
Summary
of
Subpopulation
Exposures
to
Aldrin
in
Media
Other
than
Water
.
.
.
.
.
.
.
.
.
.
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.
.
5­
16
Table
5­
6.
Summary
of
Subpopulation
Exposures
to
Dieldrin
in
Media
Other
than
Water
.
.
.
.
.
.
.
.
.
.
.
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.
5­
17
Table
6­
1.
Distribution
of
Dieldrin
in
Rats
after
104
Weeks
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
6­
7
Table
6­
2.
Relative
Tissue
Levels
of
Dieldrin
in
the
Rat
Following
a
Single
Oral
Dose
.
.
.
.
.
.
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.
.
6­
8
Table
6­
3.
Trivial
Chemical
Names
of
Aldrin,
Dieldrin
and
Their
Metabolites
.
.
.
.
.
.
.
.
6­
14
Table
9­
1.
Dose­
Response
Information
from
Key
Studies
of
Aldrin
and
Dieldrin
Toxicity
.
.
.
.
.
.
.
.
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.
.
9­
8
Table
9­
2.
Selected
Summary
Statistics
for
Occurrence
of
Aldrin
and
Dieldrin
in
Drinking
Water
.
.
.
.
.
.
.
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.
.
.
.
9­
14
Table
9­
3.
National
Population
Estimates
for
Aldrin
and
Dieldrin
Exposure
via
Drinking
Water
.
.
.
.
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.
.
.
9­
18
xi
Aldrin/
Dieldrin
 
February
2003
LIST
OF
FIGURES
Figure
2­
1.
Aldrin
Chemical
Structure
.
.
.
.
.
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.
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.
.
2­
1
Figure
2­
2.
Dieldrin
Chemical
Structure
.
.
.
.
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.
.
.
.
2­
2
Figure
3­
1.
Biodegradation
Pathways
for
Aldrin
and
Dieldrin,
With
Particular
Reference
to
Oceanic
Conditions
.
.
.
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.
3­
8
Figure
3­
2.
Photochemical
Transformations
(
Principally
Atmospheric)
Reported
for
Aldrin
and
Dieldrin
.
.
.
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.
3­
9
Figure
4­
1.
Geographic
Distribution
of
Cross­
Section
States
for
Round
2
(
SDWIS/
FED)
.
.
.
.
.
.
.
.
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.
.
.
4­
7
Figure
4­
2.
States
With
PWSs
With
Detections
of
Aldrin
for
All
States
With
Data
in
SDWIS/
FED
(
Round
2)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
.
.
.
.
4­
14
Figure
4­
3.
Round
2
Cross­
Section
States
With
PWSs
With
Detections
of
Aldrin
(
Any
PWSs
With
Results
Greater
than
the
Minimum
Reporting
Level
[
MRL];
Above)
and
Concentrations
Greater
than
the
Health
Reference
Level
(
HRL;
Below)
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
.
.
4­
15
Figure
4­
4.
Geographic
Distribution
of
Cross­
Section
States
for
Round
2
(
SDWIS/
FED)
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
4­
26
Figure
4­
5.
States
With
PWSs
With
Detections
of
Dieldrin
for
All
States
With
Data
in
SDWIS/
FED
(
Round
2)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
4­
33
Figure
4­
6.
Round
2
Cross­
Section
States
With
PWSs
With
Detections
of
Dieldrin
(
Any
PWS
With
Results
Greater
than
the
Minimum
Reporting
Level
[
MRL];
Above)
and
Concentrations
Greater
than
the
Health
Reference
Level
(
HRL;
Below)
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
4­
34
Figure
6­
1.
Distribution
Scheme
for
Dieldrin
Among
Blood
and
Various
Tissues
in
Humans
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
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.
.
.
.
.
.
6­
4
Figure
6­
2.
Metabolites
of
Aldrin
and
Dieldrin
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
6­
14
Figure
6­
3.
Proposed
Principal
Metabolic
Pathways
for
Aldrin
and
Dieldrin
.
.
.
.
.
.
.
.
.
.
6­
16
Figure
7­
1.
The
Possible
Mode
of
Action
of
Aldrin/
Dieldrin
on
Hepatocarcinogenesis
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
7­
39
1­
1
Aldrin/
Dieldrin
 
February
2003
1.0
EXECUTIVE
SUMMARY
The
U.
S.
Environmental
Protection
Agency
(
EPA)
has
prepared
this
Health
Effects
Support
Document
to
assist
in
determining
whether
to
establish
a
National
Primary
Drinking
Water
Regulation
(
NPDWR)
for
aldrin
and
dieldrin.
Case
study
reports
of
human
exposures
and
laboratory
studies
with
animals
demonstrate
that
oral
exposure
to
both
of
these
compounds
can
cause
various
adverse
systemic,
neurological,
reproductive/
developmental,
immunological,
and
genotoxic
effects.
Although
multiple
bioassays
have
established
aldrin
and
dieldrin
as
hepatocarcinogenic
in
several
strains
of
mice,
they
are
apparently
not
carcinogenic
in
rats,
and
several
large
epidemiology
studies
have
failed
to
associate
convincingly
exposure
to
them
with
cancer
in
humans.
While
some
of
these
effects
occur
only
at
moderate­
to­
high
doses,
others
have
been
observed
at
doses
lower
than
0.1
mg/
kg
bw/
day.
Nonetheless,
the
relatively
infrequent
occurrences
of
aldrin/
dieldrin
at
very
low
concentrations
indicated
by
monitoring
data,
coupled
with
the
fact
that
they
are
no
longer
manufactured
or
used
in
this
country,
indicate
that
aldrin/
dieldrin
concentrations
of
concern
are
unlikely
to
be
found
in
public
water
systems.
EPA
will
present
a
preliminary
determination
and
further
analysis
in
the
Federal
Register
Notice
covering
the
Contaminant
Candidate
List
proposals.

Chemical
Identities
and
Properties
Aldrin
(
CAS
Registry
Number
[
RN]
309­
00­
2)
is
the
most
common
name
for
the
substance
composed
of
at
least
95%
of
the
chemical
1,2,3,4,10,10­
hexachloro­
1,4,4a,
5,8,8ahexahydro
exo­
1,4­
endo­
5,8­
dimethanonaphthalene.
Technical
grade
aldrin
contains
at
least
90%
of
this
substance
(
i.
e.,
it
has
a
main
ingredient
purity
of
at
least
85.5%).
Similarly,
dieldrin
(
CAS
RN
60­
51­
1)
refers
to
the
substance
composed
of
at
least
85%
of
the
chemical
1,2,3,4,10,10­
hexachloro­
6,7­
epoxy­
1,4,4a,
5,6,7,8,8a­
octahydro­
endo­
1,4­
exo­
5,8­
dimethanonaphthalene.
Technical
grade
dieldrin
contains
at
least
95%
of
this
substance
(
i.
e.,
it
has
a
main
ingredient
purity
of
at
least
80.75%).

Dieldrin,
a
stereoisomer
of
endrin,
was
typically
produced
by
the
epoxidation
of
aldrin
with
peracetic
or
perbenzoic
acid.
In
their
"
pure"
formulations,
both
aldrin
and
dieldrin
are
composed
of
clear­
to­
white
crystals
with
densities
greater
than
water,
and
have
both
low
volatilities
and
aqueous
solubilities.
Both
are
relatively
stable
in
the
presence
of
organic
and
inorganic
alkalies
and
mild
acids,
slightly
corrosive
to
metals
upon
storage,
and
compatible
with
most
fertilizers
and
pesticides.

Aldrin/
Dieldrin
Uses,
Manufacture,
and
Environmental
Fate
Aldrin
and
dieldrin
are
synthetic
organochlorine
pesticides
that
act
as
effective
contact
and
stomach
poisons
for
insects.
Originally,
they
were
used
as
broad­
spectrum
soil
insecticides
for
the
protection
of
various
food
crops,
as
seed
dressings,
to
control
infestations
of
pests
like
ants
and
termites,
and
to
control
several
insect
vectors
of
disease.
In
1972,
the
EPA
cancelled
all
but
three
specific
uses
of
these
compounds
(
subsurface
termite
control,
dipping
of
non­
food
plant
roots
and
tops,
and
completely
contained
moth­
proofing
in
manufacturing
processes),
which
by
1987
were
voluntarily
cancelled
by
the
manufacturer.
Use
of
these
compounds
peaked
in
the
1­
2
Aldrin/
Dieldrin
 
February
2003
U.
S.
during
1966
at
19
million
lbs
for
aldrin
and
1
million
lbs
for
dieldrin.
These
compounds
have
not
been
produced
domestically
since
1974,
and
while
some
importation
of
aldrin
began
during
that
year,
this
ceased
after
1985.

Total
releases
of
aldrin/
dieldrin
to
the
environment
since
1987
are
not
known,
but
hazardous
waste
treatment
facilities
in
three
states
(
AR,
MI,
TX)
reported
releases
totaling
25,622
lbs
in
1998,
most
of
which
was
directly
to
land.
Data
from
the
Agency
for
Toxic
Substances
and
Disease
Registry
(
ATSDR)
indicate
that
these
compounds
have
been
detected
in
site
samples
from
40
different
states;
aldrin
has
been
detected
at
National
Priorities
List
(
NPL)
hazardous
waste
sites
in
31
states,
while
dieldrin
has
been
found
at
NPL
sites
in
38
states.

Under
most
environmental
conditions,
aldrin
is
largely
converted
via
biological
and/
or
abiotic
mechanisms
to
dieldrin,
which
is
significantly
more
persistent.
Most
environmental
releases
of
aldrin
and
dieldrin
are
directly
to
soil.
Because
of
low
water
solubility
and
tendency
to
bind
strongly
to
soils,
both
compounds
migrate
downward
very
slowly
through
soils
or
into
surface
or
ground
water.
Most
surface
water
aldrin/
dieldrin
has
been
attributed
to
particulate
surface
run­
off.
Over
time,
it
is
possible
that
significant
volatilization
of
aldrin/
dieldrin
might
occur,
with
subsequent
atmospheric
photodegradation
and/
or
rainfall
"
washout."
Collectively,
these
characteristics
will
foster
low
levels
of
aldrin/
dieldrin
water
contamination
over
comparatively
extended
periods
of
time.
Dieldrin's
extreme
apolarity
results
in
a
high
affinity
for
organic
matter
such
as
animal
fats
and
plant
waxes,
which
could
lead
to
its
bioaccumulation
in
the
food
chain.

Exposure
to
Aldrin/
Dieldrin
As
neither
aldrin
nor
dieldrin
has
been
used
in
the
U.
S.
since
1987,
new
releases
to
the
environment
should
not
occur.
Only
rare
exceptions
to
this
generalization
might
occur
at
hazardous
waste
treatment
facilities.
Over
time,
therefore,
the
frequency
and
magnitude
of
population
exposure
to
aldrin/
dieldrin
can
be
confidently
expected
to
decline
from
those
experienced
to
date
(
2001).
Currently
available
sampling
and
monitoring
data
suggest
that
although
potential
exposures
to
aldrin/
dieldrin
via
drinking
water
could
be
of
similar
magnitude
to
those
estimated
from
the
diet,
which
exposures
in
turn
are
likely
to
be
substantially
higher
than
those
from
breathing
air
or
ingesting
soil,
they
are
unlikely
to
occur
at
significant
frequencies
or
dose
levels.

The
data
analyzed
in
this
document
on
the
occurrence
in
drinking
water
of
aldrin/
dieldrin
were
collected
beginning
in
1993
under
"
Round
2"
of
the
Safe
Drinking
Water
Act's
Unregulated
Contaminant
Monitoring
(
UCM)
Program.
Monitoring
ended
in
January
1999,
for
small
public
water
systems
(
PWSs),
and
in
January
2001,
for
large
PWSs.
These
data,
from
34
states
and
a
number
of
Native
American
tribal
systems,
were
not
collected
utilizing
a
uniform
or
adequate
statistical
framework,
and
were
in
some
cases
incomplete
and/
or
biased.
To
partially
address
the
questionable
representativeness
of
the
combined
data
set,
a
"
national
cross­
section"
of
20
Round
2
states
(
AK,
AR,
CO,
KY,
ME,
MD,
MA,
MI,
MN,
MS,
NH,
NM,
NC,
ND,
OH,
OK,
OR,
RI,
TX,
and
WA)
was
selected.
The
procedure
used
to
construct
this
"
reasonable
representation"
of
national
occurrence
evaluated
the
individual
data
sets
for
completeness,
1­
3
Aldrin/
Dieldrin
 
February
2003
quality,
bias,
pollution
potentials
from
manufacturing/
population
density
and
from
agricultural
activity,
and
for
"
geographic
coverage"
in
relation
to
all
states.
Because
data
from
MA
were
incomplete
and
considered
abnormal
for
synthetic
organic
compounds
like
aldrin/
dieldrin
(
an
atypically
high
percentage
of
detections
in
a
relatively
small
number
of
PWSs),
Round
2
crosssection
occurrence
data
for
aldrin/
dieldrin
are
discussed
primarily
in
the
context
of
the
other
19
states.

The
data
indicate
that
each
compound
is
only
infrequently
detected
in
PWSs,
and
then,
generally,
only
at
very
low
concentrations.
With
respect
to
the
Health
Reference
Level
(
HRL,
a
preliminary
estimated
health
effect
level
used
in
these
analyses)
for
these
compounds
of
0.002
:
g/
L
(
based
on
estimated
excess
lifetime
cancer
risks
of
10
S
6),
concentrations
of
aldrin
and
dieldrin
greater
than
or
equal
to
this
level
were
detected
in
only
0.016
and
0.093%
of
the
Round
2
cross­
section
PWSs,
respectively.
These
percentages
extrapolate
nationally
to
11
PWSs
serving
38,871
people
for
aldrin,
and
61
PWSs
serving
149,827
people
for
dieldrin.
As
a
consequence
of
excluding
states
with
positively­
biased
detect
statistics,
Round
2
cross­
section
data
underestimate
the
national
occurrence
of
these
compounds
in
PWSs.
It
is
important
to
remember
that
only
one
positive
sample
(
i.
e.,
taken
at
a
single
time
point
from
a
single
sampling
location)
was
required
to
classify
a
PWS
as
one
with
aldrin
or
dieldrin
detections
 
a
practice
that
certainly
overestimates
population
exposures.

Data
from
all
the
reporting
Round
2
states
may
be
used
to
derive
more
conservative,
probably
over­
estimates
of
the
national
PWS
occurrences
of
aldrin
and
dieldrin
at
levels
$
the
HRL.
These
data
yield
respective
PWS
detection
rates
of
0.212
and
0.211%,
which
extrapolate
nationally
to
138
PWSs
serving
1,051,989
people
and
137
PWSs
serving
792,703,
respectively.
Only
five
states
(
AL,
MA,
NM,
PA,
TX)
and
eight
states
(
AL,
AR,
CT,
MA,
MD,
NC,
PA,
TX)
detected
aldrin
or
dieldrin,
respectively,
in
any
PWS.

While
the
U.
S.
Geological
Survey's
National
Ambient
Water
Quality
Assessment
(
NAWQA)
Program
did
not
analyze
for
the
presence
of
aldrin
in
ambient
ground
or
surface
waters,
it
did
analyze
for
samples
of
aquatic
biota
tissue
and
stream
bed
sediments
taken
from
591
sites
located
in
significant
watersheds
and
aquifers
from
1992
to
1995.
Aldrin
was
not
detected
in
any
of
the
aquatic
biota
samples,
but
was
detected
above
the
Method
Detection
Limit
(
MDL)
of
1
mg/
kg
at
0.4%
of
the
sites
(
detections
were
confined
to
mixed
land
use
and
agricultural
sites;
there
were
no
urban
or
forest­
rangeland
detections).
Similarly,
dieldrin
was
detected
above
the
1
mg/
kg
MDL
at
13.7%
of
the
same
sites,
as
well
as
above
the
MDL
of
5
mg/
kg
in
28.6
and
6.4%
of
whole
fish
and
bivalve
samples,
respectively.
Unlike
aldrin,
dieldrin
was
an
NAWQA
analyte
for
ambient
surface
and
ground
waters
from
1991
to
1996.
At
MDLs
of
0.001
and
0.01
mg/
L,
dieldrin
was
detected
in
4.64
and
2.39%,
respectively,
of
total
stream
surface
water
sites,
and
in
1.42
and
0.93%,
respectively,
of
total
ground
water
sites.

Relative
source
contribution
analyses
estimate
that
ratios
of
dietary
to
drinking
water
intake
range
from
1.7
to
3.8
for
aldrin,
and
from
0.9
to
8.8
for
dieldrin.
Ratios
were
computed
for
the
70
kg
adult
and
the
10
kg
child
consuming
2
L/
day
or
1
L/
day,
respectively,
of
drinking
water,
and
utilized
either
the
median
or
the
99th
percentile
concentrations
of
the
Round
2
crosssection
PWS
samples
(
detections
only)
for
aldrin
(
0.58
or
0.69
:
g/
L)
and
dieldrin
(
0.16
or
1.36
1­
4
Aldrin/
Dieldrin
 
February
2003
:
g/
L),
as
well
as
estimated
adult
and
child
total
dietary
intakes
of
aldrin
(
3.3
to
6.5
and
13
to
18
×
10!
5
mg/
kg
bw/
day,
respectively)
and
dieldrin
(
3.6
and
14
×
10!
5
mg/
kg
bw/
day,
respectively),
which
were
based
on
data
from
the
1980s
to
early­
to­
mid
1990s.

These
dietary/
drinking
water
intake
ratios
would
be
reduced
by
factors
of
approximately
3
to
6
under
the
very
conservative
approach
of
using
median
and
99th
percentile
detect
concentrations
based
on
monitoring
data
from
all
reporting
UCM
Round
2
states.
Thus,
drinking
water
appears
capable
of
potentially
providing
a
significant
portion
of
the
total
daily
dietary
intake
of
aldrin/
dieldrin
only
when
analyzed
utilizing
conservative
assumptions,
and
then
only
for
limited
populations
under
unlikely
exposure
circumstances.

Even
when
using
30­
year­
old
air
monitoring
data
that
likely
substantially
overestimate
current
daily
inhalation
intakes
of
aldrin/
dieldrin,
they
are
still
relatively
low
(
0.013
to
0.24
×
10­
5)
compared
to
dietary
estimates
and
potentially
possible
(
although
unlikely)
exposures
from
drinking
water.
Similarly,
data
available
for
dieldrin
suggest
that
ingestion
of
soil
represents
only
a
minor
exposure
pathway
for
aldrin/
dieldrin.

Toxicokinetics
of
Aldrin/
Dieldrin
Few
direct
data
were
found
in
the
literature
on
the
absorption
of
aldrin/
dieldrin,
especially
in
humans.
Dose­
related
increases
in
blood
and
adipose
tissue
levels
of
dieldrin
were
reported
in
volunteers
exposed
via
diet
to
small
amounts
for
18
to
24
months,
with
concentrations
in
the
blood
equal
to
8.6%
of
the
amount
ingested
per
day
under
steady­
state
conditions.
Inhalation
studies
using
volunteers
suggest
that
20
to
50%
of
inhaled
aldrin
vapor
may
be
absorbed
and
retained
in
the
human
body.
One
study
in
rats
estimated
that
approximately
10%
of
an
orally
administered
dose
of
aldrin
was
absorbed
via
the
gastrointestinal
tract.
Other
studies
in
rats
have
demonstrated
that
dieldrin
concentrations
in
the
blood
and
liver
increase
during
the
first
9
days
of
dietary
exposure
to
50
parts
per
million
(
ppm),
then
remain
fairly
constant
over
the
next
6
months;
also,
that
absorption
of
aldrin
and
dieldrin
is
detected
within
1
to
5
hours
after
oral
dosing
and
occurs
primarily
via
the
hepatic
portal
vein
instead
of
the
thoracic
lymph
duct.
Additionally,
uptake
of
aldrin
in
isolated,
perfused
rabbit
lungs
was
demonstrated
to
occur
in
a
biphasic
process
of
simple
diffusion.
Direct
absorption
of
aldrin/
dieldrin
through
intact
skin
has
been
reported
in
rabbits,
dogs,
monkeys,
and
humans.

Because
of
its
relatively
rapid
metabolic
conversion
to
dieldrin,
aldrin
is
infrequently
observed
in
human
tissue
and
there
is
little
information
on
its
distribution
in
human
tissue.
As
a
result
of
their
hydrophobic
nature,
the
highest
concentrations
of
aldrin/
dieldrin
and
their
metabolites
are
typically
found
in
the
adipose
tissues
of
both
humans
and
other
animals.
Based
on
several
studies
involving
volunteers
or
human
autopsies,
the
steady­
state
relative
distribution
of
dieldrin
in
whole
blood,
brain
grey
matter,
brain
white
matter,
liver,
and
adipose
tissue
is
estimated
to
be
1,
2.8,
4.2,
22.7,
and
136,
respectively.
The
leanest
individuals
appear
to
have
the
highest
adipose
tissue
concentration
of
dieldrin,
but
both
the
lowest
total
body
burden
of
dieldrin
and
the
lowest
proportion
of
total
exposure
dose
is
retained
in
their
adipose
tissue.
Blood
levels
of
dieldrin
do
not
increase
during
periods
of
surgical
stress
or
complete
fasting,
and
decline
exponentially
after
termination
of
exposure,
with
considerable
variation
among
1­
5
Aldrin/
Dieldrin
 
February
2003
individuals
(
mean
half­
lives
of
266
and
369
days
were
reported
in
2
studies).
Placental
transfer
of
dieldrin
can
occur,
resulting
in
fetal
blood
concentrations
higher
than
those
in
maternal
blood
(
1.22
vs.
0.53
mg/
kg,
respectively).

Distribution
studies
conducted
in
animals
(
rats,
mice,
guinea
pigs,
dogs,
primates,
and
various
domesticated
species)
generally
support
the
findings
from
human
studies,
at
least
qualitatively.
Exposure
to
aldrin/
dieldrin
leads
to
preferential
disposition
of
dieldrin
(
and
metabolites)
in
adipose
tissue,
with
lesser­
to­
very
small
amounts
variously
reported
in
liver,
kidney,
brain,
muscle,
lung,
blood,
and
certain
other
tissues.
In
partial
summary,
there
are
some
differences
in
distribution
parameters
among
species
and,
at
least
in
rodents,
between
sexes
(
females
reportedly
absorb
and
retain
more
dieldrin
in
their
adipose
tissue
and
most
organs
than
do
males);
blood
concentrations
appear
to
decline
more
rapidly
upon
termination
of
exposure
in
animals
than
in
humans;
redistribution
of
dieldrin
from
the
liver
to
adipose
tissue
may
occur
principally
via
the
lymphatic
system;
transplacental
transfer
of
dieldrin
has
also
been
demonstrated
in
rodents;
and
the
available
animal
data
collectively
suggest
that
distribution
patterns
of
aldrin
and
dieldrin
will
be
similar
for
most
routes
of
exposure.

As
noted
previously,
in
many
organisms
the
initial
and
principal
biotransformation
of
aldrin
following
oral
exposure
is
the
relatively
rapid,
mixed
function
oxidase­
mediated
epoxidation
to
dieldrin.
Also
referred
to
as
aldrin­
epoxidase,
these
enzymes
are
prominent
in
the
endoplasmic
reticulum
of
vertebrate
hepatocytes.
Male
rats
and
mice
appear
to
convert
more
rapidly
and
extensively
than
do
females.
In
some
extra­
hepatic
tissues
(
e.
g.,
lung)
that
contain
relatively
little
cytochrome
P­
450
activity,
in
vitro
studies
suggest
that
aldrin
may
be
epoxidized
to
dieldrin
via
an
alternate,
prostaglandin
endoperoxide
synthase
pathway,
one
which
is
dependent
on
arachidonic
acid
rather
than
on
nicotine
adenine
dinucleotide
phosphate
(
NADPH).
Additionally,
several
in
vivo
and
in
vitro
animal
studies
have
demonstrated
the
dermal
conversion
of
aldrin
to
dieldrin.
Although
data
from
humans
are
extremely
sparse,
one
excretion
study
conducted
on
workers
occupationally
exposed
to
aldrin/
dieldrin
identified
9­
hydroxy
dieldrin
as
a
fecal
metabolite.
Animal
studies
have
collectively
demonstrated
the
following
metabolites
of
dieldrin
to
be
among
the
most
significant:
pentachloroketone,
6,7­
transdihydroxydihydroaldrin
and
its
glucuronide
conjugate,
9­
hydroxy
dieldrin
and
its
glucuronide
conjugate,
and
aldrin
dicarboxylic
acid.
The
appearance
and
proportions
of
these
metabolites
can
vary
by
species,
strain,
and
sex,
as
can
the
overall
rates
of
aldrin/
dieldrin
biotransformation.

Limited
data
from
occupational
and
volunteer
studies
suggest
that
in
humans,
excretion
of
aldrin/
dieldrin
and
most
of
their
metabolites
occurs
primarily
through
the
bile
and
feces,
with
smaller
amounts
appearing
in
the
urine.
In
addition,
nursing
mothers
have
been
found
to
excrete
dieldrin
via
lactation.
Similar
findings
are
observed
in
most
animals,
although
in
rabbits
urinary
excretion
exceeds
fecal
excretion.
Again,
the
identity
and
relative
amounts
of
fecal
and
urinary
excretion
products
can
vary
somewhat
among
species
(
e.
g.,
pentachloroketone
was
identified
as
a
significant
urinary
metabolite
in
the
CFE
rat,
but
was
not
detected
in
the
CF
1
mouse),
as
well
as
between
sexes
(
biliary/
fecal
and
urinary
excretion
following
exposure
to
radiolabeled
dieldrin
was
found
to
be
higher
in
male
than
in
female
rats).
1­
6
Aldrin/
Dieldrin
 
February
2003
Adverse
Effects
from
Exposure
to
Aldrin/
Dieldrin
Data
from
the
available
literature
indicate
that
oral
exposure
to
aldrin/
dieldrin
can
induce
a
range
of
adverse
systemic,
neurological,
reproductive/
developmental,
immunological,
genotoxic,
and
tumorigenic
effects
in
humans
and/
or
animals.
Some
of
these
effects
are
manifested
only
at
moderate
to
relatively
high
doses,
but
others
have
been
observed
at
doses
lower
than
0.1
mg/
kg
bw/
day.

In
humans,
acute
exposures
to
high
concentrations
of
aldrin/
dieldrin
result
most
notably
in
toxicity
to
the
central
nervous
system;
effects
most
commonly
reported
include
hyperirritability,
convulsions,
and
coma,
sometimes
followed
by
cardiovascular
sequelae
such
as
tachycardia
and
elevated
blood
pressure.
Persistent
headache,
nausea
and/
or
vomiting,
shortterm
memory
loss,
hypothermia,
and
abnormal
electroencephalogram
patterns
have
also
been
observed.
For
adult
males,
the
acute
oral
lethal
dose
(
LD
50)
for
both
compounds
has
been
estimated
to
be
5
g,
or
about
70
mg/
kg
bw.

When
humans
have
been
exposed
for
longer
periods
to
lower
doses
of
these
compounds,
neurotoxic
symptoms
have
included
headache,
dizziness,
general
malaise,
nausea,
vomiting,
and
muscle
twitching
or
myoclonic
jerking.
In
general,
occupational
studies
indicate
that
exposure
to
aldrin/
dieldrin
does
not
result
in
adverse
hematological
or
immunological
(
e.
g.,
dermal
sensitization)
effects
in
humans.
However,
two
cases
of
immunohemolytic
anemia
have
been
linked
to
dieldrin
exposure,
as
have
several
instances
of
aplastic
anemia
to
aldrin/
dieldrin
exposure.
While
some
of
these
associations
appear
fairly
suggestive,
others
are
more
problematic.

The
available
literature
does
not
include
other
significant
adverse
health
effects
in
humans
resulting
from
longer­
term
or
chronic
exposure
to
aldrin/
dieldrin.
With
the
exception
of
several
statistically
significant
increases
in
the
incidence
of
rectal
or
liver/
biliary
cancer
that
generally
disappeared
in
follow­
up
studies,
a
variety
of
occupational/
epidemiology
studies
have
failed
to
provide
convincing
evidence
that
exposure
to
aldrin/
dieldrin
results
in
elevated
risks
of
either
cancerous
or
noncancerous
disease.
When
standardized
mortality
ratios
of
exposed
vs.
general
populations
were
computed
for
both
specific
causes
and
all
causes
of
death,
virtually
all
were
lower
than
1.0
in
both
initial
and
follow­
up
reports.

Available
animal
data
(
mouse,
rat,
guinea
pig,
rabbit,
and
dog)
indicate
oral
LD
50
values
ranging
from
33
to
95
mg/
kg
bw.
Similar
to
those
described
in
humans,
neurotoxic
effects
observed
in
animals
following
acute
to
chronic
exposure
to
aldrin/
dieldrin
include
increased
irritability,
salivation,
hyperexcitability,
tremors
followed
by
convulsions,
loss
of
body
weight,
depression,
prostrations,
and
death.
Convulsions
were
observed
in
the
rat
after
exposure
to
aldrin
for
3
days
at
10
mg/
kg
bw/
day,
as
was
brain
cell
histopathology
after
a
6­
month
exposure
to
2.75
mg/
kg
bw/
day
in
rats,
or
a
9­
month
exposure
to
0.89
mg/
kg
bw/
day
in
dogs.
Chronic
exposure
of
rats
and
mice
to
0.45
to
1.5
mg
aldrin/
kg
bw/
day
has
variously
resulted
in
hyperexcitability,
tremors,
and
clonic
convulsions.
1­
7
Aldrin/
Dieldrin
 
February
2003
Single
doses
of
0.5
to
16.7
mg
dieldrin/
kg
bw
were
reported
to
disrupt
operant
behavior
in
the
rat,
and
three
2­
to
4­
month
rat
studies
collectively
demonstrated
hyperexcitability,
tremors,
and
impaired
operant
behavior
at
Lowest­
Observed­
Adverse­
Effect
Levels
(
LOAELs)
of
2.5,
0.5,
or
0.025
mg
dieldrin/
kg
bw/
day,
respectively.
Various
long­
term
(
80
weeks
to
29
months)
rat
studies
collectively
reported
hyperexcitability,
irritability,
tremors,
and/
or
convulsions
at
LOAELs
of
0.5
to
2.5
mg
dieldrin/
kg
bw/
day.
In
another
2­
year
study
in
rats
that
had
several
potential
limitations,
cerebral
edema
and
small
degenerative
foci
were
found
at
doses
as
low
as
0.0016
mg
dieldrin/
kg
bw/
day.
In
one
2­
year
study
in
dogs,
convulsions
were
observed
at
0.5
mg
dieldrin/
kg
bw/
day,
while
another
reported
normal
electroencephalograms
at
0.05
mg
dieldrin/
kg
bw/
day.

In
a
number
of
short­
to­
intermediate
term
studies
in
rats
and
mice,
various
manifestations
of
hepatotoxicity
(
increased
relative
liver
weight,
liver
enlargement,
hepatocyte
hypertrophy,
and
elevated
DNA
synthesis;
induction
of
mixed
function
oxidases,
increased
size
and
number
of
focal
lesions
in
the
rat,
but
not
the
mouse,
following
pretreatment
with
diethyl
nitrosamine)
were
associated
with
LOAELs
ranging
from
0.5
to
1.5
mg
dieldrin/
kg
bw/
day,
and
No­
Observed­
Adverse­
Effect
Levels
(
NOAELs)
ranging
from
0.15
to
0.5
mg
dieldrin/
kg
bw/
day.
One
7­
to
10­
day
mouse
study
reported
elevated
relative
liver
weights
at
doses
as
low
as
0.015
mg
dieldrin/
kg
bw/
day
(
a
NOAEL
was
not
determined).

One
longer­
term
(
16­
month)
study
in
dogs
reported
increased
absolute
and
relative
liver
weights
and
hepatic
fatty
degeneration
at
doses
of
0.12
to
0.25
mg
aldrin/
kg
bw/
day,
but
not
0.043
to
0.091
mg
aldrin/
kg
bw/
day;
however,
no
signs
of
hepatotoxicity
were
reported
in
another
25­
month
study
in
dogs
at
0.5
mg
aldrin/
kg
bw/
day.
Liver
histopathology
was
observed
in
one
2­
year
rat
study
at
0.025
mg
aldrin/
kg
bw/
day,
as
were
enlarged
livers
at
2.5
mg
aldrin/
kg
bw/
day;
nondose­
related
liver
histopathology
was
also
seen
at
1
mg
aldrin/
kg
bw/
day,
and
increased
relative
liver
weights
at
1.5
mg/
kg
bw/
day,
in
a
second
long­
term
(
31­
month)
study
in
rats.
However,
hepatotoxicity
was
not
noted
in
several
other
long­
term
studies
in
the
mouse,
rat,
or
dog.
Similarly,
while
several
long­
term
studies
of
dieldrin
in
the
rat,
mouse,
or
dog
did
not
report
evidence
of
hepatotoxicity,
increased
absolute
and/
or
relative
liver
weights,
increased
serum
alkaline
phosphatase
activity,
and
liver
histopathology
were
collectively
observed
in
three
other
2­
year
studies
(
two
rat,
one
dog)
at
0.025
to
0.05
mg
aldrin/
kg
bw/
day.

There
are
limited
animal
data
to
suggest
that
aldrin/
dieldrin
can
induce
nephropathy
or
exacerbate
pre­
existing
nephropathy.
One
2­
year
study
in
rats
reported
that
nephritis
and
distended­
hemorrhagic
urinary
bladders
were
associated
with
a
LOAEL
of
2.5
mg
aldrin/
kg
bw/
day
and
a
NOAEL
of
0.5
mg
aldrin/
kg
bw/
day.
Exposures
to
0.043
to
0.091
mg
aldrin/
kg
bw/
day
for
up
to
16
months
were
reported
to
cause
distal
renal
tubule
vacuolation
in
female
dogs,
and
in
dogs
of
both
sexes
at
0.12
to
0.25
mg/
kg
bw/
day.
Chronic
exposure
to
5.0
and
7.5
mg
dieldrin/
kg
bw/
day
has
been
reported
to
result
in
the
development
of
hemorrhagic
and/
or
distended
urinary
bladders
in
male
rats,
usually
accompanied
by
substantial
nephritis.

In
general,
animal
studies
have
provided
only
mixed
data
that
moderate­
to­
relatively
high
doses
of
aldrin/
dieldrin
can
result
in
adverse
reproductive
or
developmental
effects.
There
are
1­
8
Aldrin/
Dieldrin
 
February
2003
some
in
vivo
and
in
vitro
data
to
suggest
that
these
compounds
may
be
weak
endocrine
disruptors,
as
various
effects
on
male
and
female
hormone
levels
and/
or
receptor
binding,
estrus
cycle,
endometrial
or
breast
cell
proliferation,
and
male
germ
cell
degeneration
and
interstitial
testicular
cell
ultrastructure
have
been
reported.
A
5­
day
exposure
of
male
mice
to
1
mg
aldrin/
kg
bw/
day
failed
to
produce
unequivocal
evidence
of
dominant
lethality,
and
a
single
exposure
of
male
mice
to
50
mg
dieldrin/
kg
bw
did
not
produce
a
significant
dominant
lethal
effect.

Among
the
effects
noted
in
several
studies
in
rats
and
dogs
at
aldrin
doses
of
0.125
to
0.3
mg/
kg
bw/
day
were
reduced
pup
survival
during
lactation,
failure
to
achieve
estrous
in
some
females,
impaired
mammary
development
and
milk
production,
and
depressed
sexual
drive
in
males;
initially,
reduced
fertility
was
also
observed
in
two
3­
generation
rat
studies
at
doses
of
0.625
to
1.38
mg
aldrin/
kg
bw/
day.

Similarly,
several
studies
using
rats,
mice,
or
dogs
have
demonstrated
that
dieldrin
doses
of
0.125
to
0.75
mg/
kg
bw/
day
can
result
in
reduced
pup
survival
during
lactation.
Dieldrin
doses
of
0.125
to
0.275
mg/
kg
bw/
day
have
also
resulted
in
initially
reduced
parental
generation
fertility
rates
in
3­
generation
rat
studies.
Another
limited
rat
study
reported
various
neural
lesions
in
pups
born
to
dams
dosed
with
as
little
as
0.004
to
0.008
mg
dieldrin/
kg
bw/
day.
Exposure
to
dieldrin
doses
of
4
mg/
kg
bw/
day
(
gestation
day
[
gd]
15
to
postpartum
day
[
ppd]
21)
or
6
mg/
kg
bw/
day
(
gd
7
to
16)
did
not
affect
fecundity,
stillbirth
or
terata
frequencies,
fetotoxicity,
or
perinatal
mortality
in
two
studies
in
rats.
However,
teratogenic
responses
(
webbed
foot,
cleft
palate,
open
eye)
were
observed
in
mice
and
hamsters
after
dieldrin
exposures
of
15
mg/
kg
bw/
day
(
gd
9)
or
30
mg/
kg
bw/
day
(
gd
7
to
9),
respectively.
Another
study
in
mice
noted
an
increase
in
supernumerary
ribs,
but
not
in
major
malformations,
after
a
dieldrin
exposure
of
3
mg/
kg
bw/
day
(
gd
7
to
16).

With
respect
to
the
immunotoxicity
of
aldrin/
dieldrin,
several
studies
in
mice
suggest
that
exposure
to
dieldrin
may
induce
immunosuppression:
single
oral
doses
of
$
18
mg/
kg
bw
have
reportedly
decreased
the
antigenic
response
to
mouse
hepatitis
virus
3;
a
10­
week
dietary
exposure
to
concentrations
as
low
as
1
ppm
(
0.15
mg/
kg
bw/
day)
increased
the
lethality
of
Plasmodium
berghei
or
Leishmania
tropica
infections;
and
3,
6,
or
18
weeks
of
dietary
exposure
to
concentrations
as
low
as
1
ppm
(
0.15
mg/
kg
bw/
day)
were
found
to
decrease
tumor
cell
killing
ability.

Numerous
long­
term
bioassays
have
convincingly
demonstrated
that
aldrin
and
dieldrin
are
hepatocarcinogens
in
several
strains
of
mice;
in
one
of
these
studies
dieldrin
was
also
judged
to
have
induced
lung,
lymphoid,
and
"
other"
tumors.
Increased
incidences
of
hepatocellular
carcinoma
and/
or
adenoma
in
mice
have
been
reported
for
doses
as
low
as
0.6
to
1.5
mg
aldrin/
kg
bw/
day
and
0.375
to
1.5
mg
dieldrin/
kg
bw/
day.
In
one
dieldrin
study,
however,
dose­
related
increases
in
the
incidence
of
hepatocellular
carcinoma
and
combined
liver
tumors,
as
well
as
decreases
in
tumor
latency,
began
at
doses
as
low
as
0.015
mg/
kg
bw/
day.
In
contrast
to
these
results,
all
of
the
available
bioassays
(
some
of
which
are
now
considered
inadequate
tests
of
carcinogenicity)
have
failed
to
demonstrate
any
evidence
of
liver
tumorigenicity
in
any
strain
of
rats
that
was
tested.
Further,
only
a
single
rat
bioassay
of
aldrin
gave
any
evidence
of
1­
9
Aldrin/
Dieldrin
 
February
2003
tumorigenicity
at
any
site
 
evidence
for
increased
incidences
of
thyroid
follicular
cell
adenoma/
carcinoma
in
males
and
females
and
adrenal
cortex
adenoma/
carcinoma
in
females,
increases
which
have
been
considered
equivocal/
suggestive
by
some,
and
unrelated
to
treatment
by
others.
As
noted
previously,
aldrin/
dieldrin's
carcinogenicity
has,
on
balance,
not
been
demonstrated
in
humans.

Much
remains
unknown
about
the
modes
of
action
that
may
underlie
the
various
toxic
effects
produced
by
exposure
to
aldrin/
dieldrin.
The
hyperexcitability
associated
with
aldrin/
dieldrin
neurotoxicity
may
arise
from
enhancement
of
synaptic
activity
throughout
the
central
nervous
system
(
CNS),
but
it
is
not
clear
whether
it
results
from
facilitated
neurotransmitter
release
at
the
nerve
terminals
or
from
reducing
the
activity
of
inhibitory
neurotransmitters
within
the
CNS.
One
hypothesis
suggests
that
dieldrin
may
act
by
inhibiting
calcium­
dependent
brain
ATPases,
which
would
inhibit
the
cellular
efflux
of
calcium
and
result
in
higher
intracellular
calcium
levels
and
subsequent
neurotransmitter
release.
Data
from
relatively
recent
studies
indicate
that
aldrin/
dieldrin's
principal
mode
of
neurotoxic
action
likely
involves
their
role
as
antagonists
of
the
membrane
receptor
for
the
inhibitory
neurotransmitter,
gamma
aminobutyric
acid
(
GABA),
and
blocking
the
influx
of
chloride
ion
through
the
GABA
A
receptor­
ionophore
complex.
Further,
an
in
vitro
study
using
fetal
rat
brain
cells
suggests
that
dieldrin
may
have
an
even
greater
functional
effect
on
dopaminergic
neurons.

From
the
available
studies,
the
carcinogenic
potential
of
aldrin/
dieldrin
appears
largely
confined
to
the
mouse,
and
it
may
not
rest
predominantly
on
genotoxicity
modes
of
action.
This
appears
most
evident
in
the
general
failure
of
aldrin/
dieldrin
to
induce
gene
point
mutations
(
28
negative
assays,
3
positive).
However,
when
considering
either
direct
DNA
damage
or
chromosome­
related
interactions
(
aberrations,
aneuploidy,
SCEs),
the
assay
results
are
significantly
more
balanced
(
15
negative,
2
most
likely
negative,
11
positive,
4
"
questionably"
positive).

Aldrin/
dieldrin's
capacity
to
inhibit
various
forms
of
in
vitro
intercellular
communication
in
both
human
and
animal
cells
may
represent
a
significant
"
epigenetic"
mode
of
carcinogenic
action
with
respect
to
their
in
vivo
effects
on
tumor
production.
Several
recent
studies
suggest
that
the
mouse­
specific
hepatocarcinogenic
effects
of
aldrin/
dieldrin
may
result
from
the
induction
of
intracellular
oxidative
stress
(
via
the
generation
of
reactive
oxygen
species
that
result
in
oxidative
damage
to
DNA,
protein,
and
lipid
macromolecules),
as
well
as
increased
hepatic
DNA
synthesis.
These
effects
generally
occur
after
aldrin/
dieldrin
treatment
in
mice,
but
not
in
rats.
After
observing
the
frequency
and
patterns
of
c­
Ha­
ras
proto­
oncogene
mutations
appearing
in
the
DNA
of
glucose­
6­
phosphatase­
deficient
hepatic
lesions
found
in
control
mice,
or
in
those
treated
with
dieldrin
or
phenobarbital,
another
study
concluded
that
the
increase
in
hepatic
lesions
(
and
thus
tumors)
resulting
from
dieldrin
treatment
principally
resulted
from
promotional,
rather
than
initiation,
events.
It
also
has
been
postulated
that
aldrin/
dieldrin
induction
of
hepatic
DNA
synthesis
may
result
from
the
modulation
of
protooncogene
expression
via
various
transcription
factors.

The
available
literature
included
almost
no
direct
evidence
for
any
human
subpopulations
that
would
be
particularly
sensitive
to
the
toxic
effects
of
aldrin/
dieldrin,
or
for
which
relevant
1­
10
Aldrin/
Dieldrin
 
February
2003
toxicokinetics
are
known
to
differ
significantly
from
those
for
the
general
population.
Speculatively,
the
fetus
and
very
young
children
might
be
at
increased
risk
from
exposures
to
aldrin/
dieldrin
as
a
result
of
immature
hepatic
detoxification
and
excretion
functions,
as
well
as
developing
target
organ
systems.
In
this
regard,
a
single
case
study
reported
that
a
3
year­
old
female
child
died
after
ingesting
approximately
8.2
mg
aldrin/
kg
bw,
which
is
roughly
an
order
of
magnitude
below
the
estimated
lethal
dose
for
adult
males.
Several
mechanistic
studies
that
describe
the
prenatal
effects
of
aldrin/
dieldrin
on
GABA
receptor
malfunctions
and
on
subsequent
behavioral
impairment
also
suggest
an
increased
sensitivity
of
children.
Declining
organ
and
immune
functions
could
potentially
render
the
elderly
more
susceptible
to
aldrin/
dieldrin
toxicity,
and
it
is
reasonable
to
expect
that
any
individuals
with
compromised
liver,
immune,
or
neurological
functions
(
as
a
result
of
disease,
genetic
predisposition
or
toxic
insult)
might
be
especially
sensitive
to
these
compounds.

Dose­
Response
Assessments
As
previously
noted,
the
acute
oral
lethal
dose
for
aldrin/
dieldrin
in
adult
humans
has
been
estimated
at
70
mg/
kg
bw,
which
is
about
3
times
the
dose
reported
to
have
induced
convulsions
within
20
minutes
of
ingestion.
Oral
LD
50
values
in
various
animal
species
for
the
two
compounds
have
been
reported
to
range
from
33
to
95
mg/
kg
bw,
and
may
be
affected
by
age
at
the
time
of
exposure.
In
rats,
LD
50
values
were
reported
at
37
mg/
kg
bw
for
young
adults,
25
mg/
kg
bw
for
2­
week­
old
pups,
and
168
mg/
kg
bw
for
newborns.

Adequate
dose­
response
relationships
have
not
been
characterized
in
humans
for
any
of
the
toxic
effects
of
aldrin/
dieldrin.
In
animals,
oral
exposure
has
produced
a
variety
of
dosedependent
systemic,
neurological,
immunological,
endocrine,
reproductive,
developmental,
genotoxic,
and
tumorigenic
effects
over
a
collective
dose
range
of
at
least
three
orders
of
magnitude
(<
0.05
to
50
mg/
kg
bw),
depending
on
endpoint
and
exposure
duration.
For
noncancer
effects,
the
U.
S.
EPA
has
determined
oral
Reference
Doses
(
RfDs)
for
both
aldrin
and
dieldrin
based
on
the
most
sensitive
relevant
toxic
effects
(
critical
effects)
reported.
For
aldrin,
the
critical
effect
was
liver
toxicity
observed
in
one
rat
study
after
chronic
exposure
to
approximately
0.025
mg/
kg
bw/
day,
the
LOAEL
and
the
lowest
dose
tested.
This
dose
was
divided
by
a
composite
uncertainty
factor
of
1,000
(
to
account
for
rat­
to­
human
extrapolation,
potentially
sensitive
human
subpopulations,
and
the
use
of
a
LOAEL
rather
than
a
NOAEL)
to
yield
an
oral
RfD
of
3
×
10­
5
mg/
kg
bw/
day.
Similarly,
for
dieldrin
a
chronic
rat
NOAEL
for
liver
toxicity
of
approximately
0.005
mg/
kg
bw/
day
was
divided
by
a
composite
uncertainty
factor
of
100
(
to
account
for
rat­
to­
human
extrapolation
and
potentially
sensitive
human
subpopulations),
yielding
an
oral
RfD
of
5
×
10­
5
mg/
kg
bw/
day.

Based
on
long­
term
mouse
bioassays,
the
EPA
has
classified
both
aldrin
and
dieldrin
as
Group
B2
carcinogens
under
the
1986
cancer
guidelines,
that
is,
as
probable
human
carcinogens
with
little
or
no
evidence
of
carcinogenicity
in
humans,
and
sufficient
evidence
in
animals.
Under
the
U.
S.
EPA's
proposed
1996/
1999
cancer
risk
assessment
guidelines,
the
weight
of
evidence
indicates
that
aldrin
and
dieldrin
could
be
classified
as
rodent
carcinogens
that
are
"
likely
to
be
carcinogenic
to
humans
by
the
oral
route
of
exposure,
but
whose
carcinogenic
potential
by
the
inhalation
and
dermal
routes
of
exposure
cannot
be
determined
because
there
1­
11
Aldrin/
Dieldrin
 
February
2003
are
inadequate
data
to
perform
an
assessment."
This
characterization
must
be
tempered
by
the
lack
of
evidence
for
significant
human
carcinogenicity
from
epidemiological
studies
and
by
the
general
lack
of
corroborative
evidence
for
carcinogenicity
in
rats.
Mechanistic
studies
suggest
that
non­
genotoxic
modes
of
action
may
underlie
or
contribute
to
aldrin/
dieldrin's
carcinogenic
potential,
but
their
relevance
to
human
carcinogenicity
is
not
fully
established,
and
a
role
for
genotoxic
mechanisms
cannot
confidently
be
eliminated
based
on
the
available
data.
Based
on
these
considerations,
the
quantitative
cancer
risk
assessments
of
aldrin
and
dieldrin
have
been
conducted
conservatively
using
the
linear­
default
model.

This
approach
has
yielded
respective
geometric
mean
cancer
potency
estimates
for
aldrin
and
dieldrin
of
17
and
16
(
mg/
kg
bw/
day)­
1.
These
result
in
drinking
water
unit
risks
of
4.9
×
10­
4
per
mg/
L
and
4.6
×
10­
4
per
mg/
L,
respectively.
For
both
compounds,
an
estimated
lifetime
excess
cancer
risk
of
10­
6
results
from
a
drinking
water
concentration
of
0.002
:
g/
L
.
This
concentration,
0.002
:
g/
L,
was
selected
as
the
Health
Reference
Level
(
HRL)
used
elsewhere
in
this
document
to
put
into
context
the
levels
of
aldrin/
dieldrin
detected
in
drinking
water.

Risk
Characterizations
and
Regulatory
Determinations
for
Aldrin/
Dieldrin
Evaluating
the
second
criterion
involves
analysis
of
public
water
system
monitoring
data,
ambient
water
concentrations
and
environmental
releases,
and
the
chemical's
environmental
fate.
Since
aldrin/
dieldrin
have
not
been
used
in
the
U.
S.
since
1987,
no
new
environmental
releases
are
expected
(
with
the
possible
exception
of
a
very
few
from
hazardous
waste
treatment
plants).
Available
data
indicate
that
these
chemicals
are
detected
very
infrequently
in
drinking
water,
and
then
at
very
low
concentrations.
Their
occurrence
in
ambient
water
appears
to
be
of
minimal
concern,
and
while
environmental
fate
data
suggest
that
they
may
continue
to
be
released
to
water
over
a
long
period
of
time,
the
concentrations
involved
will
remain
quite
low.
2­
1
Aldrin/
Dieldrin
 
February
2003
2.0
IDENTITY:
PHYSICAL
AND
CHEMICAL
PROPERTIES
Figure
2­
1.
Aldrin
Chemical
Structure
The
molecular
weight
and
chemical
formula
of
aldrin
(
CAS
RN
309­
00­
2)
are
shown
above
(
Figure
2­
1),
in
conjunction
with
two
representations
of
its
structural
formula.
Aldrin
is
the
common
name
approved
by
the
International
Standards
Organization
(
except
in
Canada,
Denmark,
and
the
former
Soviet
Union)
for
the
product
that
contains
at
least
95%
of
the
substance
identified
by
one
of
the
following
IUPAC
chemical
names
(
IARC,
1974a;
IPCS,
1989a,
b;
Lewis,
1993):

1,2,3,4,10,10­
hexachloro­
1,4,4a,
5,8,8a­
hexahydro­
exo­
1,4­
endo­
5,8­
dimethanonaphthalene;
or
(
1R,
4S,
5S,
8R)­
1,2,3,4,10,10­
hexachloro­
1,4,4a,
5,8,8a­
hexahydro­
1,4:
5,8­
dimethanonaphthalene
In
Canada,
aldrin
refers
to
the
pure
compound,
which
in
Great
Britain
is
called
HHDN.
Aldrin
has
a
significant
number
of
chemical
synonyms
and
common
trade
names
(
HSDB,
2000a;
IARC,
1974a;
IPCS,
1989a,
b;
Sittig,
1991;
USEPA,
1992),
including:

ALDOCIT
Aldrex
ALDROSOL
Compound
118
Drinox
ENT
15,949
Hexachlorohexahydro­
endo­
exo­
dimethanonaphthalene
HHDN
KORTOFIN
OCTALENE
OMS
194
SEEDRIN
2­
2
Aldrin/
Dieldrin
 
February
2003
Technical
grade
aldrin
was
formulated
to
contain
not
less
than
90%
aldrin
(
as
defined
above),
i.
e.,
not
less
than
85.5%
of
the
main
ingredient,
with
not
less
than
4.5%
insecticidal
impurities
and
not
more
than
10%
other
impurities
(
HSDB,
2000a;
IARC,
1974a;
IPCS,
1989a,
b).
Impurities
that
have
been
identified
include
a
complex
mixture
of
compounds
formed
by
the
polymerization
of
hexachlorocyclopentadiene
(
HCCPD)
and
bicycloheptadiene
(
BCH)
(
3.6
to
3.7%),
polychlorohexahydrodimethanonaphthalene
compounds
(
isodrin)
(
3.5%),
hexachlorobutadiene
(
0.5
to
0.6%),
chlordane
(
0.5%),
octachlorocyclopentene
(
0.4
to
0.5%),
toluene
(
0.3
to
0.6%),
HCCPD
(
0.2%),
HHDN
di­
adduct
(
0.1%),
BCH
(<
0.1%),
and
hexachloroethane
(<
0.1%)
(
IARC,
1974a;
IPCS,
1989a,
b).

Aldrin
has
been
formulated
into
seed
dressings
(
75%),
dust
concentrates
(
75%),
emulsifiable
concentrates
(
24
to
48%),
wettable
powders
(
20
to
40%),
granules
(
2
to
25%),
lowpercentage
dusts
(
2
to
5%),
and
mixtures
with
fertilizers
(
0.4
to
2%)
(
HSDB,
2000a;
IARC,
1974a).
Epichlorohydrin,
a
known
carcinogen,
was
sometimes
incorporated
into
the
emulsions
to
help
prevent
corrosion
by
hydrochloric
acid,
as
was
urea
into
wettable
powders
to
prevent
dehydrochlorination
by
certain
catalytically­
active
carriers
(
HSDB,
2000a).

Aldrin
is
reported
to
be
stable
in
the
presence
of
organic
and
inorganic
alkalies,
diluted
acids,
and
hydrated
metal
chlorides
(
Budavari
et
al.,
1989;
IARC,
1974a;
Lewis,
1993).
While
minimally
corrosive
to
steel,
brass,
monel,
copper,
nickel,
and
aluminum,
aldrin
can
be
slightly
corrosive
to
metals
upon
storage
as
a
result
of
the
slow
formation
of
hydrogen
chloride
(
HSDB,
2000a;
IPCS,
1989b).
Most
fertilizers,
herbicides,
fungicides,
and
insecticides
were
reported
to
be
compatible
with
aldrin
(
Lewis,
1993),
but
in
general,
contact
with
concentrated
mineral
acids,
acid
catalysts,
acid
oxidizing
agents,
phenols,
or
active
metals
should
be
avoided
(
IPCS,
1989a,
b;
Sittig,
1991).

Figure
2­
2.
Dieldrin
Chemical
Structure
Dieldrin
is
formed
by
the
epoxidation
of
aldrin
with
peracetic
or
perbenzoic
acid
(
IARC,
1974a).
Some
of
aldrin's
chemical
properties
are
summarized
later
in
Table
2­
1.
2­
3
Aldrin/
Dieldrin
 
February
2003
Table
2­
1.
Selected
Chemical­
Physical
Properties
of
Aldrin
and
Dieldrin1
Property
Aldrin
Dieldrin
Chem.
Formula
(
MW)
C
12
H
8
Cl
6
(
364.93)
C
12
H
8
Cl
6
O
(
380.93)

Physical
State
Clear
to
white
crystals;
tan
to
dark
brown
solid
(
technical)
Clear
to
white
crystals;
buff
to
light
tan
flakes
(
technical)

Melting
Point
104!
105.5
°
C;
49!
60
°
C
(
technical)
175!
177
°
C;
>
95
°
C
(
technical)

Boiling
Point
145
°
C
(
at
2
mm
Hg)
330
°
C
Density
(
at
20
°
C)
1.6!
1.7
g/
cc;
1.54
g/
cc
(
technical)
1.75
g/
cc;
1.62
g/
cc
(
technical)

Solubility
(
Water)
0.027
mg/
L
(
at
27
°
C);
also
reported
as
0.20
mg/
L
(
at
25
°
C)
0.1!
0.195
mg/
L
(
at
20!
29
°
C)

Solubility
(
Organic
Solvents)
Moderately
to
very
sol.
in
most
paraf­
finic
and
aromatic
hydrocarbons,
esters,
ketones,
and
halogenated
solvents,
less
so
in
alcohols;
>
600
g/
L
in
acetone,
benzene,
and
xylene
(
at
27
°
C)
Moderately
sol.
in
common
organic
solvents,
except
aliphatic
petroleum
hydrocarbons,
and
methanol
(
in
g/
L
at
20
°
C:
400
!
benzene,
220
!
acetone,
10
!
methanol)

Log
K
ow
3.01
or
6.50;
7.4
(
technical)
5.40;
6.2
(
technical)

Log
K
oc
4.96
3.87
Vapor
Pressure
(
20
°
C)
2.3!
7.5
x
10!
5
mm
Hg
3.1
x
10
!
6
or
1.78
x
10!
7
mm
Hg
Vapor
Pressure
(
25
°
C)
1.4
x
10!
4
mm
Hg
or
6
x
10!
6
mm
Hg
5.89
x
10!
6,
7.78
x
10!
7,
or
1.8
x
10!
7
mm
Hg
Henry's
Law
Constant
(
at
25
°
C)
3.2
x
10!
4
atm­
m3/
mol
or
1.27
x
10!
5
atm­
m3/
mol
(
est.)
5.8
x
10!
5
atm­
m3/
mol
or
1.51
x
10!
5
atm­
m3/
mol
Odor
Mild
chemical
odor
Mild
chemical
odor
Odor
Threshold
0.017
mg/
L
(
water)
0.3
mg/
m3
(
air)
0.04
mg/
L
(
water)
NA
(
air)

Conversion
Factors2
(
at
25
°
C,
1
atm)
1
ppm
=
14.96
mg/
m3
(
at
25
°
C,
1
atm)
1
ppm
=
15.61
mg/
m3
(
at
25
°
C,
1
atm)

1
ATSDR
(
2000);
Budavari
et
al.
(
1989);
HSDB
(
2000a,
b);
IARC
(
1974a,
b);
IPCS
(
1989b);
Lewis
(
1993);
Sittig
(
1991);
Verschueren
(
1983).
2
ATSDR
(
2000).
2­
4
Aldrin/
Dieldrin
 
February
2003
The
molecular
weight
and
chemical
formula
of
dieldrin
(
CAS
RN
60­
57­
1)
are
shown
above
(
Figure
2­
2),
in
conjunction
with
two
representations
of
its
structural
formula.
Dieldrin
is
the
common
name
approved
by
the
International
Standards
Organization
(
except
in
Canada,
Denmark,
and
the
former
Soviet
Union)
for
the
product
that
contains
at
least
85%
of
the
substance
identified
by
one
of
the
following
IUPAC
chemical
names
(
IARC,
1974b;
IPCS,
1989a,
b;
Lewis,
1993):

1,2,3,4,10,10­
hexachloro­
6,7­
epoxy­
1,4,4a,
5,6,7,8,8a­
octahydro­
endo­
1,4­
exo­
5,8­
dimethanonaphthalene;
or
(
1R,
4S,
5S,
8R)­
1,2,3,4,10,10­
hexachloro­
1,4,4a,
5,6,7,8,8a­
octahydro­
6,7­
epoxy­
1,4:
5,8­
dimethanonaphthalene
In
Canada,
dieldrin
refers
to
the
pure
compound,
which
in
Great
Britain
is
called
HEOD.
Dieldrin
has
a
significant
number
of
chemical
synonyms
and
common
trade
names
(
HSDB,
2000b;
IARC,
1974b;
IPCS,
1989a,
b;
Sittig,
1991;
USEPA,
1988),
including:

ALVIT
Compound
497
DIELDREX
DIELMOTH
ENT
16,225
HEOD
Hexachloroexpoxyoctahydro­
endo­
exo­
dimethanonaphthalene
Illoxol
Octalux
OMS
18
QUINTOX
Red
Shield
TERMITOX
Technical
grade
dieldrin
was
formulated
to
contain
not
less
than
95%
dieldrin
(
as
defined
above),
i.
e.,
not
less
than
80.75%
of
the
main
ingredient;
however,
it
was
available
in
the
United
States
in
a
formulation
containing
100%
active
ingredient,
i.
e.,
not
less
than
85%
HEOD,
with
not
less
than
15%
related
insecticidally­
active
compounds
(
HSDB,
2000b;
IARC,
1974a;
IPCS,
1989a,
b;
Lewis,
1993).
Impurities
reportedly
found
in
technical
grade
dieldrin
include
aldrin,
other
polychloroepoxyoctahydrodimethanonaphthalenes
(
including
endrin,
3.5%),
free
HCl
(<
0.4%),
and
water
(<
0.1%)
(
HSDB,
2000b;
IARC,
1974b;
IPCS,
1989a,
b).

Dieldrin
has
been
formulated
into
wettable
powders
(
40
to
75%),
oil
solutions
(
18
to
20%),
emulsifiable
concentrates
(
15
to
20%),
granules
(
5%),
seed
dressings,
dusts,
and
mixtures
with
fertilizers
(
HSDB,
2000b;
IARC,
1974b).

Dieldrin
is
reported
to
be
stable
in
the
presence
of
organic
and
inorganic
alkalies,
mild
acids
commonly
used
in
agriculture,
and
light
(
Budavari
et
al.,
1989;
IARC,
1974b;
IPCS,
1989a,
b),
although
it
may
react
with
sunlight
to
produce
photodieldrin
(
IARC,
1974b).
As
with
2­
5
Aldrin/
Dieldrin
 
February
2003
aldrin,
dieldrin
can
be
slightly
corrosive
to
metals
upon
storage
as
a
result
of
the
slow
formation
of
hydrogen
chloride
(
HSDB,
2000b;
IPCS,
1989b).
Most
fertilizers,
herbicides,
fungicides,
and
insecticides
were
reported
to
be
compatible
with
dieldrin
(
Lewis,
1993),
but
in
general,
contact
with
concentrated
mineral
acids,
acid
catalysts,
acid
oxidizing
agents,
phenols,
or
active
metals
(
iron,
copper,
sodium)
should
be
avoided
(
Budavari
et
al.,
1989;
IPCS,
1989a,
b;
Sittig,
1991).
Dieldrin
is
formed
by
the
epoxidation
of
aldrin
with
peracetic
or
perbenzoic
acid
(
IARC,
1974a,
b),
and
is
a
stereoisomer
of
endrin
(
Budavari
et
al.,
1989).
It
reportedly
reacts
with
hydrogen
bromide
to
give
the
bromohydrin
(
HSDB,
2000b).
Some
of
dieldrin's
chemical
properties
are
summarized
in
Table
2­
1.
2­
6
Aldrin/
Dieldrin
 
February
2003
References
ATSDR.
2000.
Agency
for
Toxic
Substances
and
Disease
Registry.
Toxicological
profile
for
aldrin/
dieldrin
(
Update).
Draft
for
public
comment.
Atlanta,
GA:
US
Dept.
of
Health
and
Human
Services,
Public
Health
Service,
ATSDR.

Budavari,
S,
M.
J.
O'Neil,
A.
Smith,
and
P.
E.
Heckelman
(
eds.).
The
Merck
index,
11th
ed.
Rahway,
NJ:
Merck
&
Co.,
Inc.,
pp.
223,
490.

HSDB.
2000a.
Hazardous
Substances
Data
Bank.
Aldrin.
Retrieved
Sep.
20,
2000.
Bethesda,
MD:
National
Library
of
Medicine,
Specialized
Information
Services
Division,
Toxicology
and
Environmental
Health
Information
Program,
TOXNET.

HSDB.
2000b.
Hazardous
Substances
Data
Bank.
Dieldrin.
Retrieved
Sep.
20,
2000.
Bethesda,
MD:
National
Library
of
Medicine,
Specialized
Information
Services
Division,
Toxicology
and
Environmental
Health
Information
Program,
TOXNET.

IARC.
1974a.
International
Agency
for
Research
on
Cancer.
Evaluation
of
the
carcinogenic
risk
of
chemicals
to
humans.
Aldrin.
Lyon,
France:
IARC
Monograph
5:
25­
38.

IARC.
1974b.
International
Agency
for
Research
on
Cancer.
Evaluation
of
the
carcinogenic
risk
of
chemicals
to
humans.
Dieldrin.
Lyon,
France:
IARC
Monograph
5:
125­
156.

IPCS.
1989a.
International
Programme
on
Chemical
Safety.
Aldrin
and
dieldrin
health
and
safety
guide.
Health
and
safety
guide
no.
21.
Geneva,
Switzerland:
World
Health
Organization,
IPCS.

IPCS.
1989b.
International
Programme
on
Chemical
Safety.
Aldrin
and
dieldrin.
Environmental
health
criteria
91.
Geneva,
Switzerland:
World
Health
Organization,
IPCS.

Lewis,
Sr.,
R.
J.
1993.
Hawley's
condensed
chemical
dictionary,
12th
ed.
New
York,
NY:
Van
Nostrand
Reinhold
Company,
pp.
32,
387.

Sittig,
M.
1991.
Handbook
of
toxic
and
hazardous
chemicals
and
carcinogens,
3rd
ed.,
vol.
1.
Park
Ridge,
NJ:
Noyes
Publications,
pp.
6­
64,
598­
601.

USEPA.
1992.
US
Environmental
Protection
Agency.
Aldrin
drinking
water
health
advisory.
Washington,
DC:
USEPA
Office
of
Water.

USEPA.
1988.
US
Environmental
Protection
Agency.
Dieldrin
health
advisory.
Washington,
DC:
USEPA
Office
of
Water.

Verschueren,
K.
1983.
Handbook
of
environmental
data
on
organic
chemicals,
2nd
ed.
New
York,
NY:
Van
Nostrand
Reinhold,
pp.
168­
173,
513­
518.
3­
1
Aldrin/
Dieldrin
 
February
2003
3.0
USES
AND
ENVIRONMENTAL
FATE
This
section
summarizes
information
derived
from
cited
secondary
references
pertaining
to
the
uses,
manufacture,
and
environmental
fate
of
aldrin
and
dieldrin.

3.1
Uses
and
Manufacture
These
compounds
are
organochlorine
pesticides
that
act
as
highly
effective
contact
and
stomach
poisons
for
insects
(
IPCS,
1989a).
Aldrin
was
used
as
a
broad­
spectrum
soil
insecticide
(
generally
at
0.5
to
5
kg/
hectare)
for
the
protection
of
corn,
potato,
citrus,
and
other
crops
against
termites,
corn
rootworms,
seed
corn
beetles
and
maggots,
wireworms,
rice
water
weevil,
grasshoppers,
Japanese
beetles,
etc.,
as
well
as
a
seed
dressing
for
rice
and
to
combat
ant
and
termite
infestations
of
wooden
structures
(
ATSDR,
2000;
IPCS,
1989a,
b;
USEPA,
1992).
Dieldrin
was
once
used
similarly
in
agriculture,
but
no
longer;
it
was
then
used
principally
to
protect
wooden
structures
against
ant
and
termite
attack,
in
industry
for
protection
against
termites,
wood
borers
and
textile
pests,
and
as
a
residual
spray
and
larvacide
for
the
control
of
several
insect
vectors
of
disease
(
ATSDR,
2000;
IPCS,
1989a,
b;
USEPA,
1988).

The
US
Department
of
Agriculture
banned
all
uses
of
aldrin
and
dieldrin
in
1970,
but
in
1972
under
the
authority
of
the
Federal
Insecticide,
Fungicide
and
Rodenticide
Act
(
FIFRA),
the
EPA
permitted
their
use
in
three
cases:
subsurface
ground
insertion
for
termite
control,
dipping
of
non­
food
plant
roots
and
tops,
and
mothproofing
of
woolen
textiles
and
carpets
under
conditions
of
no
effluent
discharge
(
ATSDR,
2000;
USEPA,
1980).
The
latter
two
registered
uses
were
abandoned
by
the
manufacturer
in
1974,
as
was
the
ground­
insertion
termiticide
use
in
1987;
therefore,
all
uses
of
aldrin
and
dieldrin
have
been
canceled
(
ATSDR,
2000;
USEPA,
1980).

In
the
United
States,
the
use
of
aldrin
peaked
at
19,000,000
lbs
in
1966
and
had
declined
to
about
10,500,000
lbs
by
1970;
concurrently,
dieldrin
use
declined
from
1,000,000
lbs
to
about
650,000
lbs
(
USEPA,
1980).
There
was
some
importation
of
these
compounds
during
the
1970s
and
early­
mid
1980s;
the
USEPA
has
reported
that
no
aldrin
has
been
imported
since
1985
(
ATSDR,
2000).
Aldrin
was
not
imported
into
the
United
States
prior
to
the
1974
cancellation
decision;
however,
Shell
International
(
Holland)
imported
the
chemical
for
limited
use
from
1974
to
1985
(
with
the
exception
of
1979
and
1980,
when
imports
were
temporarily
suspended).
An
estimated
1
to
1.5
million
lbs
of
aldrin
were
imported
annually
from
1981
to
1985,
after
which
time
importation
ceased.
By
1987,
all
uses
of
aldrin
had
been
cancelled
voluntarily
by
the
manufacturer
(
ATSDR,
2000).
In
1972,
USEPA
cancelled
all
but
the
following
three
uses
of
dieldrin:
subsurface
ground
insertion
for
termite
control,
the
dipping
of
non­
food
plant
roots
and
tops,
and
mothproofing
in
manufacturing
processes
using
completely
closed
systems.
This
cancellation
decision
was
finalized
in
1974.
By
1987,
all
uses
of
dieldrin
had
been
cancelled
voluntarily
by
its
manufacturer
(
the
Shell
Chemical
Company)
(
ATSDR,
2000).
3­
2
Aldrin/
Dieldrin
 
February
2003
3.2
Environmental
Release
and
Fate
Aldrin
is
listed
as
a
Toxic
Release
Inventory
(
TRI)
chemical.
In
1986,
the
Emergency
Planning
and
Community
Right­
to­
Know
Act
(
EPCRA)
established
the
Toxic
Release
Inventory
(
TRI)
of
hazardous
chemicals.
Created
under
the
Superfund
Amendments
and
Reauthorization
Act
(
SARA)
of
1986,
EPCRA
is
also
sometimes
known
as
SARA
Title
III.
The
EPCRA
mandates
that
larger
facilities
publicly
report
when
TRI
chemicals
are
released
into
the
environment.
This
public
reporting
is
required
for
facilities
with
more
than
10
full­
time
employees
that
annually
manufacture
or
produce
more
than
25,000
pounds,
or
use
more
than
10,000
pounds,
of
a
TRI
chemical
(
USEPA,
1996/
1999;
USEPA,
2000a).

Under
these
conditions,
facilities
are
required
to
report
the
pounds
per
year
of
aldrin
released
into
the
environment
both
on­
and
off­
site.
The
production,
import,
and
use
of
aldrin
had
been
cancelled
by
the
time
the
TRI
was
instated;
therefore,
no
release
or
transfer
data
were
reported.
In
1995,
Resource
Conservation
and
Recovery
Act
(
RCRA)
Subtitle
C
hazardous
waste
treatment
and
disposal
facilities
were
added
to
the
list
of
those
facilities
required
to
present
release
data
to
the
TRI.
This
addition
became
effective
for
the
1998
reporting
year,
which
is
the
most
recent
TRI
data
currently
available.
Waste
treatment
facilities
from
three
states
(
AR,
MI,
TX)
reported
releases
of
aldrin
in
1998,
with
on­
and
off­
site
releases
totaling
25,622
pounds.
The
on­
site
quantity
is
subdivided
into
air
emissions,
surface
water
discharges,
underground
injections,
and
releases
to
land.
Most
of
the
aldrin
released
to
the
environment
was
released
directly
to
land
(
22,000
lbs)
(
USEPA,
2000b).

Although
the
TRI
data
can
be
useful
in
giving
a
general
idea
of
release
trends,
it
is
far
from
exhaustive
and
has
significant
limitations.
For
example,
only
industries
that
meet
TRI
criteria
(
at
least
10
full­
time
employees
and
the
manufacture
and
processing
of
quantities
exceeding
25,000
lbs/
year,
or
use
of
more
than
10,000
lbs/
year)
are
required
to
report
releases.
These
reporting
criteria
do
not
account
for
releases
from
smaller
industries.
Also,
the
TRI
data
is
meant
to
reflect
releases
and
should
not
be
used
to
estimate
general
exposure
to
a
chemical
(
USEPA,
2000c).

Aldrin
is
included
in
the
Agency
for
Toxic
Substances
and
Disease
Registry's
(
ATSDR)
Hazardous
Substance
Release
and
Health
Effects
Database
(
HazDat).
This
database
records
detections
of
listed
chemicals
in
site
samples;
aldrin
was
detected
in
40
states
(
states
without
detections
are
AZ,
DE,
HI,
ME,
MS,
MT,
NV,
NM,
OR,
WY)
(
ATSDR,
2000).
The
National
Priorities
List
(
NPL)
of
hazardous
waste
sites,
created
in
1980
by
the
Comprehensive
Environmental
Response,
Compensation
&
Liability
Act
(
CERCLA),
is
a
listing
of
some
of
the
most
health­
threatening
waste
sites
in
the
United
States.
Aldrin
was
detected
in
NPL
hazardous
waste
sites
in
31
states
(
USEPA,
1999).

Dieldrin
is
also
included
in
the
ATSDR's
HazDat.
Dieldrin
was
detected
in
40
states
(
states
without
detections
are
AZ,
DE,
HI,
MN,
MT,
NV,
NM,
OR,
UT,
WY)
(
ATSDR,
2000).
Dieldrin
was
detected
in
NPL
hazardous
waste
sites
in
38
states
(
USEPA,
1999).
3­
3
Aldrin/
Dieldrin
 
February
2003
In
summary,
aldrin
and
dieldrin
have
not
been
produced
in
the
United
States
since
1974,
and
all
uses
of
the
pesticide
were
cancelled
by
1987.
Aldrin
had
been
used
mostly
on
corn
and
citrus
products.
Dieldrin
had
been
used
mostly
on
corn,
potatoes,
tomatoes,
and
citrus
products.
Aldrin
was
imported
to
the
United
States
from
Holland
from
1974
to
1985
(
with
the
exception
of
1979
and
1980)
in
quantities
of
approximately
1
to
1.5
million
lbs/
year.
TRI
data
from
1998
suggest
that
aldrin
continues
to
be
released
into
the
environment,
even
though
the
chemical
is
no
longer
produced
or
used
in
the
United
States.
Aldrin's
presence
and
persistence
in
the
environment
is
evidenced
by
detections
of
the
compound
in
hazardous
waste
sites
in
at
least
31
states
(
at
NPL
sites),
as
well
as
detections
in
site
samples
in
at
least
40
states
(
listed
in
ATSDR's
HazDat).

Most
aldrin
introduced
into
the
environment
is
relatively
rapidly
converted
through
epoxidation
to
dieldrin,
which
in
turn
is
notably
persistent
in
the
environment
due
to
its
very
low
solubility
in
water
and
its
extremely
low
volatility.
Because
dieldrin
is
also
extremely
apolar,
it
displays
a
high
affinity
for
fat
and
is
thus
retained
in
animal
fats,
plant
waxes,
and
other
similar
organic
matter
in
the
environment.
This
fat
solubility
can
lead
to
a
progressive
accumulation
of
dieldrin
in
the
food
chain,
which
theoretically
could
eventually
produce
concentrations
in
organisms
that
might
exceed
lethal
limits
to
predators
or
consumers
(
Sittig,
1991;
USEPA,
1980).

Environmental
Media
Transport
and
Distribution
Given
the
historical
uses
of
aldrin
and
dieldrin,
their
point
of
entry
into
the
environment
has
most
typically
been
the
soil
(
IPCS,
1989b).
Because
of
their
strong
adsorption
to
soils
and
their
low
aqueous
solubilities,
significant
downward
leaching
of
these
compounds
through
the
soil
profile
would
not
be
anticipated
(
ATSDR,
2000;
HSDB,
2000a,
b;
IPCS,
1989b).
As
discussed
further
below,
most
aldrin
in
the
soil
is
gradually
converted
to
dieldrin
under
most
environmental
conditions
(
ATSDR,
2000;
HSDB,
2000a;
IPCS,
1989b).
Field
studies
of
the
application
of
aldrin
to
the
surface
layer
of
various
types
of
soils
have
demonstrated
nearly
quantitative
adsorption
by
organic
matter
and
clay
minerals,
and
that
even
5
years
after
application,
residual
aldrin
and
dieldrin
were
still
found
in
the
surface
layer
with
very
little
penetration
to
lower
soil
depths
(
IPCS,
1989b).
Water
has
been
found
to
compete
with
aldrin
for
adsorption
sites
in
clay
minerals,
and
thus
aldrin
binds
to
a
greater
extent
when
the
soil
is
dry;
in
dry
soils,
mineral
components
play
the
largest
role
in
adsorption,
whereas
in
moist
soils,
organic
materials
are
predominant;
and
other
factors
being
equal,
adsorption
is
expected
to
be
the
lowest
in
sandy
soils
having
minimal
organic
content
(
IPCS,
1989b).

In
one
summarized
study,
aldrin
was
applied
to
the
upper
5
inches
of
a
silt
loam
soil
(
HSDB,
2000a).
Combining
the
results
for
non­
disked
soil
with
those
of
soil
disked
for
one
summer
only,
the
reported
distribution
of
residual
aldrin
after
10
years
by
soil
depth
was
as
follows:
11
to
13%
(
0
to
2
inches),
29
to
33%
(
2
to
4
inches),
29
to
33%
(
4
to
6
inches),
23
to
29%
(
6
to
9
inches).
In
a
study
by
Weisgerber
et
al.
(
1974),
aldrin
was
quantified
at
different
soil
depths
3
to
6
months
after
its
application
at
about
3
kg/
ha
to
soils
used
for
growing
corn
in
several
countries.
Their
findings
are
summarized
below
in
Table
3­
1.
As
is
readily
apparent,
3­
4
Aldrin/
Dieldrin
 
February
2003
aldrin
demonstrated
little
proclivity
to
migrate
down
through
the
various
soil
profiles;
similar
results
were
observed
for
soils
in
England
and
Germany
used
to
grow
wheat
(
Weisgerber
et
al.,

Table
3­
1.
Aldrin
Mobility
in
Soils
Used
to
Grow
Corn1
Soil
Depth
(
cm)
Residual
Aldrin
Levels
in
Soils
Used
to
Grow
Corn2:
ppm
(%
Total
Extractable)

Germany
England
Spain
United
States
0!
10
0.78
(
78%)
1.30
(~
100%)
0.83
(
96.5%)
0.50
(
98%)

10!
20
0.18
(
18%)
<
0.01
(<
1%)
0.02
(
2.3%)
0.01
(
1.96%)

20!
40
0.03
(
3%)
<
0.01
(<
1%)
0.01
(
1.2%)
<
0.01
(<
1%)

40!
60
<
0.01
(<
1%)
<
0.01
(<
1%)
<
0.01
(<
1%)
<
0.01
(<
1%)

1
From
Weisgerber
et
al.
(
1974).
2
Measured
5,
6,
4,
or
3
months
(
respectively
by
country)
after
the
application
of
about
3
kg
aldrin/
ha.

1974),
and
for
various
laboratory
studies
of
soil
samples
in
columns
that
were
eluted
with
water
(
HSDB,
2000a;
IPCS,
1989b).

In
a
laboratory
test
of
six
types
of
soil
placed
in
chromatographic
columns,
the
percentage
of
applied
dieldrin
that
eluted
with
1600
ml
of
water
varied
from
1%
in
loam
soil,
to
65%
in
soil
containing
93%
sand
(
IPCS,
1989b).
Little
dieldrin
leaching
was
observed
in
a
similar
column
experiment
involving
3
soil
types
eluted
with
about
30
L
of
water
over
120
hours
(
IPCS,
1989b),
and
even
with
high
temperatures
and
prolonged
leaching,
dieldrin
has
been
considered
essentially
immobile
(
HSDB,
2000b).
Experimentally
determined
log
soil
sorption
coefficients
(
K
oc)
of
2.61
to
4.45
for
aldrin
and
3.87
for
dieldrin
further
suggest
that
these
compounds
are
not
highly
mobile
in
soils
and
will
not
appreciably
leach
to
groundwater
(
HSDB,
2000a,
b).
In
areas
with
poorly
controlled
erosion,
surface
run­
off
can
carry
particle­
associated
aldrin
and
dieldrin
into
surface
waters;
in
the
absence
of
sediment,
however,
rain
water
run­
off
does
not
appear
to
be
a
major
transport
mechanism
(
ATSDR,
2000;
HSDB,
2000a,
b;
IPCS,
1989b).
The
equilibrium
ratio
of
dieldrin
concentration
in
soil
to
that
in
water
was
shown
to
be
100
to
500
for
mineral
soils
and
likely
to
be
5
to
6
times
higher
for
aldrin
(
IPCS,
1989b).
Vapor
diffusion
is,
generally,
regarded
as
the
principal
mechanism
whereby
aldrin
and
dieldrin
ascend
the
soil
profile.
The
role
of
upward
mass
flow
in
capillary
water
through
a
moisture
gradient,
though
demonstrated
in
laboratory
studies,
is
now
thought
to
be
relatively
insignificant
in
the
field
(
IPCS,
1989b).

Most
studies
have
concluded
that
the
observed,
relatively
rapid
loss
of
aldrin
and
dieldrin
from
soil
during
the
first
few
months
after
application
is
principally
attributable
to
volatilization
processes
(
ATSDR,
2000;
IPCS,
1989b).
There
is
substantial
evidence
for
this.
Mosquitoes
were
shown
to
be
killed
by
vapors
emanating
from
treated
soils
and
it
is
known
that
when
aldrin
3­
5
Aldrin/
Dieldrin
 
February
2003
is
incorporated
into
soil,
it
is
most
readily
lost
from
the
surface
layer
(
IPCS,
1989b).
Various
laboratory
studies
have
reportedly
(
ATSDR,
2000;
HSDB,
2000a,
b;
IPCS,
1989b)
demonstrated
that:
volatilization
of
aldrin
is
significantly
faster
than
that
of
dieldrin
(
about
20­
fold,
in
one
case);
chamber
rates
of
volatilization
for
each
chemical
decrease
with
time
(
about
50%
over
6
to
7
hours
in
one
experiment
with
dieldrin);
volatilization
of
aldrin
from
sands
increases
(
from
trace
levels
to
up
to
7.33%
after
6
hours)
with
increased
water
content
in
the
sands
and/
or
increased
humidity
in
the
air
passing
over
the
sands;
and
volatilization
rates
of
aldrin
from
sand,
loam,
and
humus
during
the
first
or
second
hour
after
application
were
1.08,
0.21,
and
0.08,
or
0.59,
0.18,
and
0.09%
per
ml
evaporated
water,
respectively.

Actual
field
studies
on
volatilization
losses
from
soil
are
limited
in
number
and
appear
available
only
for
dieldrin
(
ATSDR,
2000;
IPCS,
1989b).
Reported
volatilization
losses
include
2.8%
after
18
weeks
and
4.5%
after
1
year.
In
one
study
involving
a
very
high
application
rate
(
22
kg/
ha
or
10
ppm)
to
soils
under
three
different
soil
moisture
conditions,
volatilization
losses
after
5
months
were
18%
in
a
plot
kept
moist
by
irrigation,
7%
in
a
non­
irrigated
plot
receiving
only
natural
rainfall,
and
only
2%
in
a
plot
flooded
to
a
depth
of
10
cm.

Related
studies
examining
the
overall
loss
(
by
any
mechanism)
of
aldrin
or
dieldrin
from
soil
have
been
reviewed
(
ATSDR,
2000;
HSDB,
2000a;
IPCS,
1989b;
Verscheuren,
1983).
After
several
years
of
field
application
of
aldrin
at
three
different
rates,
residues
were
shown
to
be
higher
in
clay
loam
than
in
sandy
loam
soils
(
half­
lives
of
79
to
97
vs.
36
to
45
days,
respectively),
although
the
rate
of
conversion
to
dieldrin
was
higher
in
the
latter
(
ATSDR,
1993;
HSDB,
2000a).
An
early
study
examined
various
Illinois
soils
that
had
been
treated
with
aldrin,
demonstrating
that
aldrin
was
indeed
transformed
to
dieldrin,
and
concluding
that
loss
of
related
residues
was
a
two­
stage
process
 
a
comparatively
rapid
phase
during
the
first
year
after
application
in
which,
typically,
~
75%
of
the
applied
dose
was
lost.
An
extended
second
phase
displayed
residue
half­
lives
of
2
to
4
years,
perhaps
due
to
increased
content
of
the
more
stable
dieldrin
in
the
total
residue
(
IPCS,
1989b).
This
same
qualitative
result
was
observed
when
aldrin
was
applied
to
muck
and
loam
soils,
with
respective
half­
lives
of
3.75
and
2.40
months
during
the
first
half
year
and
then
13.0
and
9.7
months
for
the
following
3
years
(
HSDB,
2000b).

Following
the
application
of
1.5
kg
aldrin/
ha
to
flooded
soil,
approximately
56,
45,
26,
12,
and
0%
remained
after
30,
90,
120,
240,
and
270
days,
respectively
(
HSDB,
2000b).
Similarly,
3.5
years
after
the
application
of
20
or
200
lbs
of
aldrin/"
6
inch"
acre
to
a
Miami
silt
loam,
only
1.12
and
2.55%
remained,
respectively
(
HSDB,
2000b).
Other
reported
studies
have
demonstrated
an
increase
in
aldrin
loss
from
soils
with
increasing
temperature,
more
rapid
loss
under
upland
(
80%
water­
saturated)
than
under
flooded
conditions,
and
more
rapid
loss
from
the
upper
layers
of
most
soils
(
HSDB,
2000b).
Although
some
contrary
findings
have
been
reported,
aldrin
losses
from
temperate
soils
often
appear
more
rapid
than
from
tropical
soils
(
IPCS,
1989b).
Separate
studies
carried
out
with
dieldrin
suggest
residue
rate
losses
that
are
considerably
slower
than
those
observed
for
aldrin,
but
the
reported
range
is
wide
(
IPCS,
1989b);
one
study
reported
an
average
time
of
8
years
for
the
disappearance
of
95%
of
the
dieldrin
residues;
however,
much
slower,
as
well
as
intermediate,
rates
can
also
be
found
in
the
literature.
Verschueren
(
1983)
indicates
a
period
of
1
to
6
years
for
the
disappearance
of
75
to
100%
of
3­
6
Aldrin/
Dieldrin
 
February
2003
aldrin
from
soils;
for
dieldrin,
comparable
indicated
values
were
3
to
25
years
for
75
to
100%,
and
12.8
years
for
95%.

As
noted
previously,
aldrin
and
dieldrin
are
highly
resistant
to
being
leached
from
soils,
and
as
a
consequence,
they
have
only
rarely
been
observed
in
ground
water
samples
(
ATSDR,
2000;
IPCS,
1989b).
By
contrast,
surface
waters
have
frequently
been
reported
to
contain
small
amounts
of
these
pesticides
(
more
frequently
dieldrin),
probably
as
a
result
of
surface
run­
off
of
rain
water
in
which
most
of
the
residues
are
adsorbed
to
sediments
(
ATSDR,
2000;
HSDB,
2000a,
b;
IPCS,
1989b).
The
ultimate
fate
of
these
small
residue
amounts
is
not
known
with
certainty,
but
adsorption
to
sediments,
volatilization,
and
bioconcentration
have
been
postulated
to
play
the
most
significant
roles,
with
certain
degradation
mechanisms
(
especially
abiotic)
also
involved
to
some
extent
(
ATSDR,
2000;
HSDB,
2000a,
b;
IPCS,
1989b).

While
volatilization
is
considered
an
important
pathway
for
water
residues
of
these
compounds,
conflicting
data
are
reported
in
the
literature
(
e.
g.,
volatilization
half­
life
for
dieldrin
of
hours
to
months)
(
HSDB,
2000b).
Rates
are
expected
to
vary
directly
with
wind
and
water
current
velocities,
and
inversely
with
the
depth
of
the
water
body
(
HSDB,
2000a).
Half­
lives
for
the
volatilization
of
aldrin
from
pure
water
and
from
three
natural
waters
were
reported
to
be
0.38
and
0.59
to
0.60
hours,
respectively.
From
a
different
study,
volatilization
rates
from
water
during
the
first
and
second
hours
were
reported
to
be
16.3
and
6.03%
per
ml
evaporated
water,
respectively
(
HSDB,
2000a).
Verschueren
(
1983)
and
HSDB
(
2000a)
indicate
a
derived
half­
life
value
of
185
hours
(
7.7
days)
for
aldrin
in
a
1m
column
of
water
at
25
°
C.
Using
a
water
solubility
of
0.20
mg/
L
and
a
vapor
pressure
of
6
×
10!
6
mm
Hg
(
both
measured
at
25
°
C),
an
estimated
Henry's
Law
constant
of
1.27
×
10!
5,
and
reasonable
assumptions
for
wind
velocity,
current
velocity
and
water
depth,
half­
lives
for
aldrin
in
streams,
rivers,
and
lakes
were
calculated
as
105.5
hours,
133.9
hours,
and
6873.1
hours
(
286.4
days),
respectively
(
HSDB,
2000a).
For
dieldrin,
Verschueren
(
1983)
and
IPCS
(
1989b)
indicate
a
derived
half­
life
value
of
12,940
hours
(
539.2
days)
in
a
1
m
column
of
water
at
25
°
C.
A
cited
experimental
volatilization
rate
for
dieldrin
in
water
is
5%
of
the
reaeration
rate,
which,
using
typical
reaeration
rates
for
ponds,
rivers,
and
lakes,
yields
estimated
evaporation
half­
lives
for
dieldrin
of
72,
14,
and
52
days,
respectively;
however,
values
as
short
as
6
to
9
hours
under
certain
laboratory
conditions
have
been
reported
(
HSDB,
2000b).

From
the
previous
discussion,
it
is
apparent
that
a
substantial
portion
of
the
aldrin
and
dieldrin
used
in
agriculture
is,
generally,
considered
to
reach
the
atmosphere
(
ATSDR,
2000;
HSDB,
2000a,
b;
IPCS,
1989b).
Although
there
are
data
to
suggest
that
dieldrin
may
be
transported
great
distances
in
the
atmosphere,
in
general,
only
small
amounts
have
been
detected
by
global
atmospheric
sampling
(
ATSDR,
2000;
IPCS,
1989b).
Washout
by
rain
may
play
an
important
role
in
preventing
atmospheric
accumulation
of
these
compounds,
but
the
significance
of
this
mechanism
is
called
into
question
by
observations
of
no
detectable
levels
of
aldrin
or
dieldrin
in
soils
adjacent
to
treated
areas
(
IPCS,
1989b).
As
further
discussed
below,
various
atmospheric
degradation
mechanisms
may
also
play
a
key
role
in
minimizing
accumulation.
3­
7
Aldrin/
Dieldrin
 
February
2003
Environmental
Degradation
The
principal
transformation
of
aldrin
that
occurs
in
all
aerobic
and
biologically
active
soils
is
its
epoxidation
to
dieldrin
(
ATSDR,
2000;
HSDB,
2000a;
IPCS,
1989b).
This
reaction
has
also
been
observed
in
plants,
but
does
not
occur
under
anaerobic
conditions
(
IPCS,
1989b).
Soil
transformation
to
aldrin
dicarboxylic
acid
has
also
been
well
established
(
ATSDR,
2000;
IPCS,
1989b).
Fungi
and
other
soil
microbes
have
been
demonstrated
to
degrade
aldrin
in
culture
(
ATSDR,
2000;
HSDB,
2000a;
IPCS,
1989b).
Dieldrin
is
much
more
resistant
to
biodegradation
than
aldrin,
and
thus
microbial
degradation
is
likely
only
a
minor
pathway
for
the
loss
of
dieldrin
from
soils
(
ATSDR,
2000;
HSDB,
2000b;
IPCS,
1989b).
This
is
reflected
in
the
long
times
(
years)
that
have
been
reported
for
dieldrin
half­
lives
or
times
required
for
50
to
100%
loss
(
see
previous
discussion).
There
is
some
evidence
that
certain
microbes
can
metabolize
dieldrin
to
photodieldrin
and
that
this
is
more
likely
to
occur
under
anaerobic
conditions.
A
number
of
studies
have
detected
low
to
very
low
soil
concentrations
of
photodieldrin
(
ATSDR,
2000;
HSDB,
2000b;
IPCS,
1989b).
Although
not
biodegraded
in
standard
screening
tests,
a
number
of
soil
microorganisms
have
been
isolated
that
are
capable
of
degrading
dieldrin
to
limited
degrees
(
ATSDR,
2000;
HSDB,
2000b;
IPCS,
1989b).

Under
aqueous
conditions,
biodegradation
of
aldrin
is
expected
to
be
slow;
none
was
observed
through
the
third
subculture
with
one
mixed
culture
inoculum
from
sewage,
while
an
activated
sludge
biodegraded
1.5%
of
an
initial
amount
of
aldrin
over
an
unspecified
amount
of
time
(
HSDB,
2000a).
A
water
surface
film
collected
off
the
coast
of
Hawaii
degraded
8.1%
of
added
aldrin
to
its
diol
after
30
days;
a
pure
culture
of
a
marine
alga
degraded
23.3%
of
the
initial
aldrin
to
dieldrin
and
5.2%
to
the
diol;
and
a
pure
culture
of
Aerobacter
aerogenes
was
reported
to
degrade
36
to
46%
of
an
initial
amount
of
aldrin
within
24
hours
(
HSDB,
2000a).
Under
anaerobic
aqueous
conditions,
aldrin
is
not
epoxidized
to
dieldrin,
but
has
been
reported
to
be
completely
degraded
to
other
compounds
within
60
days
by
an
anaerobic
sewage
sludge
(
ATSDR,
2000;
IPCS,
1989b).
Although
no
biodegradation
of
dieldrin
was
reported
in
some
studies
of
river
waters,
microorganisms
isolated
from
certain
lake
water
and
lake­
bottom
sediments
may
be
able
to
transform
some
dieldrin
to
photodieldrin
under
anaerobic
conditions
(
ATSDR,
2000).

However,
dieldrin
was
not
significantly
degraded
under
anaerobic
conditions
by
an
active
waste
water
sludge
or
by
sewage
sludge
microorganisms
in
2
studies
and
was
only
degraded
by
11%
after
48
hours
or
by
24%
after
32
days
in
2
other
studies
(
ATSDR,
2000).
By
comparison,
an
aerobic
activated
sludge
was
able
to
degrade
55%
of
the
initial
level
of
dieldrin
in
9
days,
another
activated
sludge
achieved
dieldrin
degradation
of
30
to
60%
(
time
frame
not
specified
in
review),
and
a
mixed
anaerobic
microbial
culture
degraded
10
:
g
dieldrin/
ml
by
50%
in
30
days
(
ATSDR,
2000).
Some
biodegradation
pathways
of
aldrin
and
dieldrin
are
illustrated
below
in
Figure
3­
1,
taken
from
Verschueren
(
1983).
Although
intended
to
describe
metabolism
under
oceanic
conditions,
they
are
relevant
to
other
soil
and
fresh
water
environments
as
previously
discussed.

Various
abiotic
processes
may
also
contribute
to
the
environmental
degradation
of
aldrin
and
dieldrin,
although
their
role
seems
generally
to
be
considered
relatively
limited
(
ATSDR,
3­
8
Aldrin/
Dieldrin
 
February
2003
2000;
HSDB,
2000a,
b;
IPCS,
1989b).
The
high
reactivity
of
hydroxyl
and
other
atmospheric
free
radicals
could
possibly
play
a
role
in
the
degradation
of
aldrin
and
dieldrin
occurring
as
vapors
(
IPCS,
1989b)
and
the
half­
life
for
vapor
phase
aldrin
reacting
with
photochemically
generated
hydroxyl
radicals
has
been
estimated
at
35
minutes
(
HSDB,
2000a).
As
might
be
expected
from
its
weak
absorption
to
wavelengths
above
290
nm,
sunlamp
photolysis
of
aldrin
vapor
has
been
observed
to
be
rather
slow
 
60%
in
1
week,
vs.
16%
in
a
dark
control
(
HSDB,
2000a).
Both
aldrin
and
dieldrin
are
susceptible
to
photochemical
reactions
following
irradiation
by
sunlight
or
Figure
3­
1.
Biodegradation
Pathways
for
Aldrin
and
Dieldrin,
With
Particular
Reference
to
Oceanic
Conditions
(
Verschueren,
1983)

UV
under
abiotic
laboratory
conditions,
with
epoxidation
and
isomerization
transformations
resulting
in
the
formation
of
photoaldrin
and
photodieldrin
(
ATSDR,
2000;
HSDB,
2000b;
IPCS,
1989b).
These
reactions
are
illustrated
below
in
Figure
3­
2,
taken
from
Verschueren
(
1983).
Photodieldrin
is
believed
to
be
a
stable
photoproduct
of
aldrin
as
it
no
longer
contains
a
chromophore.
It
has,
in
fact,
proven
resistant
to
further
photolysis
(
ATSDR,
2000).

Other
experimental
work
found
that
while
photoaldrin
was
produced
upon
sunlight
or
ultraviolet
light
(
UV)
irradiation
of
aldrin,
the
major
photoproduct
was
an
unbridged
compound
that
had
lost
a
chlorine
atom
from
the
3
position;
the
yield
of
photoaldrin
(
and
photodieldrin
from
dieldrin)
was
also
found
to
be
substantially
enhanced
in
the
presence
of
benzophenone
or
other
ketones
(
IPCS,
1989b).
Photoproducts
arising
from
the
loss
of
chlorine
atoms
have
also
been
observed
upon
the
irradiation
of
photoaldrin
and
photodieldrin
in
the
presence
of
triethylamine
(
ATSDR,
2000).
Based
on
reactions
with
hydroxyl
radicals,
the
atmospheric
half­
life
of
dieldrin
has
been
estimated
at
approximately
1
day,
but
could
be
longer
if
it
is
associated
with
particulate
matter
(
ATSDR,
2000).
Again,
it
should
be
noted
that
while
small
amounts
of
dieldrin
have
3­
9
Aldrin/
Dieldrin
 
February
2003
been
found
in
some
atmospheric
samples,
neither
aldrin,
photoaldrin,
nor
photodieldrin
has
been
detected
(
ATSDR,
2000;
IPCS,
1989b).
Therefore,
if
the
latter
two
photoproducts
occur
to
any
significant
extent
in
the
atmosphere,
they
do
not
appear
stable
enough
to
accumulate.

When
irradiated
with
UV
or
natural
sunlight
in
an
oxygenated
aqueous
solution,
aldrin
underwent
little
degradation
unless
amino
and
humic
acids
commonly
found
in
natural
waters
were
also
present
(
ATSDR,
2000;
IPCS,
1989b).
Photolysis
half­
lives
of
4.7
to
11
days
for
thin
Figure
3­
2.
Photochemical
Transformations
(
Principally
Atmospheric)
Reported
for
Aldrin
and
Dieldrin
(
Verschueren,
1983)

films
of
aldrin
irradiated
at
>
300
nm
have
been
reported
and
exposure
of
an
aldrin
film
to
sunlight
for
1
month
resulted
in
a
solution
containing
2.6%
aldrin,
9.6%
photoaldrin,
4.1%
dieldrin,
24.1%
photodieldrin,
and
59.7%
of
an
unidentified
photoproduct
(
HSDB,
2000a).
The
persistence
of
aldrin
in
river
water
was
studied
in
sealed
glass
jars
that
were
maintained
under
sunlight
and
artificial
fluorescent
light
conditions;
amounts
remaining
after
1
hour,
1
week,
2
weeks,
4
weeks,
and
8
weeks
were
100,
100,
80,
40,
and
20%,
respectively
(
Verschueren,
1983;
HSDB,
2000a).
The
conversion
was
principally
to
dieldrin
(
Verschueren,
1983).
Irradiation
at
238
nm
for
48
hours
converted
75%
of
the
aldrin
in
filtered
natural
field
water
to
dieldrin
(
ATSDR,
2000).

Hydrolysis
is
not
a
significant
abiotic
degradation
mechanism
for
aqueous
dieldrin,
as
it
occurs
with
a
half­
life
of
>
4
years;
however,
aqueous
dieldrin
will
reportedly
degrade
to
photodieldrin
in
the
presence
of
sunlight
with
an
approximate
half­
life
of
2
to
4
months
with
the
process
being
accelerated
in
waters
containing
photosensitizers
(
HSDB,
2000b).
In
somewhat
contrary
findings,
when
the
persistence
of
dieldrin
was
studied
in
sealed
glass
jars
of
river
water
that
were
maintained
under
sunlight
and
artificial
fluorescent
light
conditions,
100%
of
the
initial
dieldrin
was
reported
to
be
still
present
after
8
weeks
(
Verschueren,
1983).
3­
10
Aldrin/
Dieldrin
 
February
2003
While
it
is
possible
that
some
aldrin
and
dieldrin
may
undergo
photochemical
degradation
(
as
a
result
of
UV
irradiation
in
surface
layers),
only
small
amounts
of
photodieldrin
have
been
observed
in
soil
samples,
and
the
extent
to
which
these
may
have
resulted
from
microbial
action
is
not
certain
(
ATSDR,
2000;
IPCS,
1989b).
It
appears
that
photochemical
reactions
may
be
responsible
for
the
epoxidation
of
some
aldrin
to
dieldrin,
and
some
dieldrin
to
photodieldrin,
that
has
been
observed
on
the
leaf
surfaces
of
various
plants
(
IPCS,
1989b).

With
respect
to
other
abiotic
mechanisms,
dieldrin
has
been
reported
to
be
susceptible
to
ozone­
mediated
degradation,
and
the
clay
diluents
used
in
dust
formulations
of
aldrin
and
dieldrin
(
especially
acidic
kaolinite
and
attapulgite)
have
been
reported
to
contribute
to
their
decomposition
(
IPCS,
1989b).

Bioaccumulation
As
suggested
by
their
relatively
high
K
ow
s,
both
aldrin
and
dieldrin
have
moderate
to
high
potentials
for
bioaccumulation
(
ATSDR,
2000;
HSDB,
2000a,
b;
IPCS,
1989b).
Aldrin
and
dieldrin
uptake
by
plants
has
been
reported
to
be
substantially
higher
in
root
crops
than
in
grain
crops;
root
crops
(
e.
g.,
carrots,
radishes,
and
turnips)
are
much
more
likely
to
take
up
residues
from
treated
soils,
whereas
it
is
rare
in
grain
crops
for
residues
to
reach
detectable
levels
in
the
grain
(
IPCS,
1989b).
In
one
model
ecosystem
study,
corn
was
planted
in
vermiculite
soil
to
which
2.09
ppm
radiolabeled
aldrin
had
been
applied;
after
14
days,
the
corn
contained
2.83
ppm
radiolabeled
residue,
of
which
0.762
ppm
was
aldrin
and
1.538
ppm
dieldrin
(
ATSDR,
2000).
About
78%
of
the
residues
were
found
in
the
roots,
with
the
remainder
in
the
shoots.
The
mechanism
of
uptake
into
plants
for
these
compounds
is
not
clear.
It
may
vary
considerably
with
species
and
the
nature
of
the
soils
in
which
they
are
grown,
and
apparently
involves
both
absorption
through
roots
and
absorption
of
vapors
through
leaves
(
ATSDR,
2000;
IPCS,
1989b).
A
vole
was
introduced
into
this
same
model
ecosystem
on
day
15,
and
after
5
days
was
found
to
have
aldrin
and
dieldrin
concentrations
of
0.08
and
3.56
ppm,
respectively
(
ATSDR,
2000).

The
bioaccumulation
and
biomagnification
of
aldrin
occur
mostly
through
its
conversion
products
(
IPCS,
1989b).
Biotransfer
factors
(
BTFs)
for
beef
and
milk,
defined
as
the
ratio
of
a
compound
in
beef
or
milk
(
mg/
kg)
to
its
daily
intake
by
the
animal
(
mg/
day),
have
been
estimated
for
aldrin
to
be
0.085
and
0.023,
respectively
(
ATSDR,
2000).
In
vegetables,
a
bioconcentration
factor
(
BCF,
the
ratio
of
a
compound's
concentration
in
above
ground
plant
parts
to
that
in
soil)
of
0.021
has
been
calculated
for
aldrin
(
ATSDR,
2000).
Similarly,
BTFs
for
beef
and
milk
and
a
vegetable
BCF
have
been
estimated
for
dieldrin,
these
being
0.008,
0.011,
and
0.098,
respectively
(
ATSDR,
2000).
BCFs
for
these
compounds
in
various
aquatic
organisms
(
fish,
molluscs,
algae,
waterflea,
etc.)
have
been
reported
to
be
in
the
range
of
100
to
15,000,
while
in
various
amphibian,
avian,
earthworm,
and
mammalian
species
values
have
been
of
the
order
of
2
to
400
BCFs
(
HSDB,
2000a,
b;
IPCS,
1989b;
Verschueren,
1983).

Environmental
Fate
Summary
In
summary,
aldrin
that
is
applied
to
soil
can
be
expected
to
largely
be
converted
to
dieldrin
through
both
biological
and
abiotic
mechanisms.
Dieldrin
is
much
more
persistent
and
3­
11
Aldrin/
Dieldrin
 
February
2003
both
compounds
will
strongly
adsorb
to
sediment
or
dust
particles.
Potential
for
leaching
into
ground
water
is
low,
but
soil
run­
off
of
rain
water
may
carry
particle­
adsorbed
residues
into
surface
waters.
Substantial
volatilization
of
both
compounds
to
the
atmosphere
is
thought
to
occur,
where
significant
levels
of
photochemical
epoxidation,
isomerization,
and
reaction
with
free
radicals
(
hydroxyl
radical)
may
take
place.
Washout
of
atomospheric
aldrin
and
dieldrin
may
also
be
significant.
Monitoring
data
suggest
that
dieldrin
is
widely
dispersed
in
the
atmosphere.
However,
while
the
ultimate
fate
of
it
and
its
related
photoproducts
remains
unclear,
it
appears
they
do
not
accumulate
in
the
atmosphere.
Biodegradation
of
aldrin
is
generally
slow
and
along
with
hydrolysis,
is
thought
to
be
an
unimportant
fate
process
for
dieldrin.
Bioconcentration
and
bioaccumulation
of
these
compounds
and
their
residues
are
significant
and,
in
addition
to
their
being
continuing
contaminants
of
soil,
water,
and
air,
they
are
often
found
in
aquatic
organisms,
wildlife,
foods,
and
humans
(
HSDB,
2000a,
b;
IPCS,
1989b;
USEPA,
1980).
3­
12
Aldrin/
Dieldrin
 
February
2003
References
ATSDR.
2000.
Agency
for
Toxic
Substances
and
Disease
Registry.
Toxicological
profile
for
aldrin/
dieldrin
(
Update).
Draft
for
public
comment.
Atlanta,
GA:
U.
S.
Dept.
of
Health
and
Human
Services,
Public
Health
Service,
ATSDR.

ATSDR.
1993.
Agency
for
Toxic
Substances
and
Disease
Registry.
Toxicological
profile
for
aldrin/
dieldrin:
Update.
Atlanta,
GA:
U.
S.
Dept.
of
Health
and
Human
Services,
Public
Health
Service,
ATSDR.

HSDB.
2000a.
Hazardous
Substances
Data
Bank.
Aldrin.
Retrieved
Sep.
20,
2000.
Bethesda,
MD:
National
Library
of
Medicine,
Specialized
Information
Services
Division,
Toxicology
and
Environmental
Health
Information
Program,
TOXNET.

HSDB.
2000b.
Hazardous
Substances
Data
Bank.
Dieldrin.
Retrieved
Sep.
20,
2000.
Bethesda,
MD:
National
Library
of
Medicine,
Specialized
Information
Services
Division,
Toxicology
and
Environmental
Health
Information
Program,
TOXNET.

IPCS.
1989a.
International
Programme
on
Chemical
Safety.
Aldrin
and
dieldrin
health
and
safety
guide.
Health
and
safety
guide
no.
21.
Geneva,
Switzerland:
World
Health
Organization,
IPCS.

IPCS.
1989b.
International
Programme
on
Chemical
Safety.
Aldrin
and
dieldrin.
Environmental
health
criteria
91.
Geneva,
Switzerland:
World
Health
Organization,
IPCS.

Sittig,
M.
1991.
Handbook
of
toxic
and
hazardous
chemicals
and
carcinogens,
3rd
ed.,
vol.
1.
Park
Ridge,
NJ:
Noyes
Publications,
pp.
6­
64,
598­
601.

USEPA.
2000a.
What
is
the
Toxic
Release
Inventory?
Available
on
the
Internet
at:
http://
www.
epa.
gov/
tri/
general.
htm
Last
modified
February
28,
2000.

USEPA.
2000b.
TRI
Explorer:
Geographic
Report.
Available
on
the
Internet
at:
http://
www.
epa.
gov/
triexplorer/
geography.
htm.
Last
modified
May
5,
2000.

USEPA.
2000c.
The
Toxic
Release
Inventory
(
TRI)
and
Factors
to
Consider
when
Using
TRI
Data.
Available
on
the
Internet
at:
http://
www.
epa.
gov/
tri/
tri98/
98over.
pdf.
Last
modified
August
11,
2000.
Link
to
site
at:
http://
www.
epa.
gov/
tri/
tri98.

USEPA.
1999.
Superfund
Hazardous
Waste
Site
Basic
Query
Form.
Available
on
the
Internet
at:
http://
www.
epa.
gov/
superfund/
sites/
query/
basic.
htm.
Last
modified
December
1,1999.

USEPA.
1996/
1999.
U.
S.
Environmental
Protection
Agency.
Proposed
Cancer
Guidelines.
Available
on
the
Internet
at
http://
www.
epa.
gov/
ORD/
WebPubs/
carcinogen
/
carcin.
pdf.
USEPA.
1992.
U.
S.
Environmental
Protection
Agency.
Aldrin
drinking
water
health
advisory.
Washington,
DC:
USEPA
Office
of
Water.
3­
13
Aldrin/
Dieldrin
 
February
2003
USEPA.
1988.
U.
S.
Environmental
Protection
Agency.
Dieldrin
health
advisory.
Washington,
DC:
USEPA
Office
of
Water.

USEPA.
1980.
U.
S.
Environmental
Protection
Agency.
Ambient
water
quality
criteria
for
aldrin/
dieldrin.
Document
no.
EPA
440/
5­
80­
019.
Washington,
DC:
USEPA
Office
of
Water,
Office
of
Water
Regulations
and
Standards,
Criteria
and
Standards
Division.

Verschueren,
K.
1983.
Handbook
of
environmental
data
on
organic
chemicals,
2nd
ed.
New
York,
NY:
Van
Nostrand
Reinhold,
pp.
168­
173,
513­
518.

Weisgerber,
I.,
J.
Kohli,
R.
Kaul,
W.
Klein,
and
F.
Korte.
1974.
Fate
of
aldrin­
14C
in
maize,
wheat
and
soils
under
outdoor
conditions.
J.
Agric.
Food
Chem.
22(
4):
609­
612
(
as
cited
in
HSDB,
2000a;
IPCS,
1989b).
4­
1
Aldrin/
Dieldrin
 
February
2003
4.0
EXPOSURE
FROM
DRINKING
WATER
4.1
Aldrin
4.1.1
Ambient
Occurrence
To
understand
the
presence
of
a
chemical
in
the
environment,
an
examination
of
ambient
occurrence
is
useful.
In
a
drinking
water
context,
ambient
water
is
source
water
existing
in
surface
waters
and
aquifers
before
treatment.
The
most
comprehensive
and
nationally
representative
data
describing
ambient
water
quality
in
the
United
States
are
being
produced
through
the
United
States
Geological
Survey's
(
USGS)
National
Water
Quality
Assessment
(
NAWQA)
program.
(
NAWQA,
however,
is
a
relatively
young
program
and
complete
national
data
are
not
yet
available
from
their
entire
array
of
sites
across
the
nation.)

Data
Sources
and
Methods
The
USGS
instituted
the
NAWQA
program
in
1991
to
examine
water
quality
status
and
trends
in
the
United
States.
NAWQA
is
designed
and
implemented
in
such
a
manner
as
to
allow
consistency
and
comparison
between
representative
study
basins
located
around
the
country,
facilitating
interpretation
of
natural
and
anthropogenic
factors
affecting
water
quality
(
Leahy
and
Thompson,
1994).

The
NAWQA
program
consists
of
59
significant
watersheds
and
aquifers
referred
to
as
"
study
units."
The
study
units
represent
approximately
two­
thirds
of
the
overall
water
usage
in
the
United
States
and
a
similar
proportion
of
the
population
served
by
public
water
systems.
Approximately
one­
half
of
the
nation's
land
area
is
represented
(
Leahy
and
Thompson,
1994).

To
facilitate
management
and
make
the
program
cost
effective,
approximately
one­
third
of
the
study
units
at
a
time
engage
in
intensive
assessment
for
a
period
of
3
to
5
years.
This
is
followed
by
a
period
of
less
intensive
research
and
monitoring
that
lasts
between
5
and
7
years.
This
way
all
59
study
units
rotate
through
intensive
assessment
over
a
10­
year
period
(
Leahy
and
Thompson,
1994).
The
first
round
of
intensive
monitoring
(
1991
to
1996)
targeted
20
watersheds.
This
first
group
was
more
heavily
slanted
toward
agricultural
basins.
A
national
synthesis
of
results
from
these
study
units
focusing
on
pesticides
and
nutrients
has
been
compiled
and
analyzed
(
Kolpin
et
al.,
1998;
Larson
et
al.,
1999;
USGS,
1999a).

Aldrin
was
not
an
analyte
for
either
the
ground
water
or
the
surface
water
NAWQA
studies
included
in
the
pesticide
and
nutrient
national
synthesis
(
Kolpin
et
al.,
1998;
Larson
et
al.,
1999;
USGS,
1999b).
Because
of
analytical
and
budget
constraints
the
NAWQA
program
targets
certain
pesticides,
many
of
which
have
high
use
and/
or
have
potential
environmental
significance
(
Larson
et
al.,
1999;
USGS,
1999a).
Aldrin
may
have
been
excluded
because
it
has
not
been
used
in
agriculture
since
the
early
1970s
and
all
of
its
uses
were
discontinued
in
the
mid­
1980s
(
USGS,
1999a).
Also,
aldrin
breaks
down
in
the
environment
to
dieldrin
(
among
other
degradates),
a
compound
that
was
analyzed
in
the
NAWQA
studies
(
USGS,
1999b).
Finally,
aldrin
persisting
in
the
environment
is
more
likely
to
be
found
in
sediments
or
biotic
tissues
4­
2
Aldrin/
Dieldrin
 
February
2003
because
of
its
strong
hydrophobicity
and
sorption
potential
(
ATSDR,
1993;
Nowell,
1999;
USGS,
2000).
Consequently,
NAWQA
investigators
focused
their
aldrin
occurrence
studies
on
bed
sediments
and
aquatic
biota
tissue
(
Nowell,
1999).

Aldrin
is
an
organochlorine
insecticide.
As
a
group,
organochlorines
are
hydrophobic
and
resist
degradation.
Hydrophobic
("
water
hating")
compounds
have
low
water
solubilities
and
strong
tendencies
to
sorb
to
organic
material
in
sediments
and
accumulate
in
the
tissue
of
aquatic
biota,
where
they
can
persist
for
long
periods
of
time
(
ATSDR,
1993;
USGS,
2000).
Organochlorines
may
be
present
in
bed
sediments
and
tissues
of
aquatic
systems
even
when
they
are
undetectable
in
the
water
column
using
conventional
methods
(
Nowell,
1999).

To
determine
their
presence
in
hydrologic
systems
of
the
United
States,
the
NAWQA
program
has
investigated
organochlorine
pesticide
detections
in
bed
sediments
and
biotic
tissue,
focusing
on
the
organochlorine
insecticides
that
were
used
heavily
in
the
past
(
Nowell,
1999).
The
occurrence
of
aldrin,
one
of
the
top
three
insecticides
used
for
agriculture
in
the
1960s
and
widely
used
to
kill
termites
in
structures
until
the
mid
1980s,
was
investigated
in
this
study
(
Nowell,
1999;
USGS,
1999a).
Sampling
was
conducted
at
591
sites
from
1992
to
1995
in
the
20
NAWQA
study
units
where
the
first
round
of
intensive
assessment
took
place.
Two
of
these
basins,
the
Central
Nebraska
Basins
and
the
White
River
Basin
in
Indiana,
are
located
in
the
corn
belt
where
aldrin
use
was
heavy
during
the
1960s.
Details
regarding
sampling
techniques
and
analytical
methods
are
described
by
Nowell
(
1999).

Results
Aldrin
was
not
detected
in
aquatic
biota
tissue
samples.
However,
it
was
detected
in
stream
bed
sediment
samples.
The
occurrence
frequencies
above
the
Method
Detection
Limit
(
MDL)
of
1
µ
g/
kg
and
basic
summary
statistics
indicate
that
occurrence
in
sediments
is
very
low
(
Table
4­
1).
Both
the
median
and
95th
percentile
concentrations
were
reported
as
non­
detections
(<
MDL)
across
all
land
use
categories.

Aldrin
was
detected
in
stream
bed
sediments
only
at
agricultural
or
mixed
land
use
sites,
perhaps
reflecting
the
heavy
agricultural
use
in
the
late
1960s
and
early
1970s.
Interesting,
in
light
of
the
more
recent
termiticide
use,
no
urban
detections
were
reported.
This
may
be
partly
a
function
of
the
NAWQA
sampling
design
that
targeted
basins
more
representative
of
agricultural
and
mixed
land
use
conditions
for
the
first
round
of
intensive
monitoring
from
which
these
sediment
data
were
produced
(
see
Section
4.1.1.1).
Data
from
later
rounds
are
not
yet
available.
The
occurrence
of
a
toxic
compound
in
stream
sediments
is
pertinent
to
drinking
water
concerns
because
some
desorption
of
the
compound
from
sediments
into
water
will
occur
through
equilibrium
reactions,
although
in
very
low
concentrations.
The
occurrence
of
aldrin
in
sediments
is
also
quite
low
(
see
Table
4­
1).
4­
3
Aldrin/
Dieldrin
 
February
2003
Table
4­
1.
Aldrin
Detections
in
Stream
Bed
Sediments1
Detection
Frequency
(%
Samples
>
MDL
of
1
µ
g/
kg)
Concentration
Percentiles
(
All
Samples;
µ
g/
kg
Dry
Weight)

Median
95th
Maximum
urban
0.0%
nd2
nd
nd
mixed
0.5%
nd
nd
3
agricultural
0.6%
nd
nd
2.2
forest­
rangeland
0.0%
nd
nd
nd
all
sites
0.4%
nd
nd
3
1
Nowell,
1999.

2
Not
detected
in
concentration
greater
than
MDL.

4.1.2
Drinking
Water
Occurrence
The
Safe
Drinking
Water
Act
(
SDWA),
as
amended
in
1986,
required
Public
Water
Systems
(
PWSs)
to
monitor
for
specified
"
unregulated"
contaminants,
on
a
5­
year
cycle,
and
to
report
the
monitoring
results
to
the
states.
Unregulated
contaminants
do
not
have
an
established
or
proposed
National
Primary
Drinking
Water
Regulation
(
NPDWR);
however,
they
are
contaminants
that
were
formally
listed
and
required
for
monitoring
under
federal
regulations.
The
intent
was
to
gather
scientific
information
on
the
occurrence
of
these
contaminants
to
enable
a
decision
as
to
whether
or
not
regulations
were
needed.
All
non­
purchased
community
water
systems
(
CWSs)
and
non­
purchased
non­
transient
non­
community
water
systems
(
NTNCWSs),
with
greater
than
150
service
connections,
were
required
to
conduct
this
unregulated
contaminant
monitoring.
Smaller
systems
were
not
required
to
conduct
this
monitoring
under
federal
regulations,
but
were
required
to
be
available
to
monitor
if
the
state
decided
such
monitoring
was
necessary.
Many
states
collected
data
from
smaller
systems.
Additional
contaminants
were
added
to
the
Unregulated
Contaminant
Monitoring
(
UCM)
program
in
1991
(
USEPA,
1991)
for
required
monitoring
that
began
in
1993
(
USEPA,
1992).

Aldrin
has
been
monitored
under
the
SDWA
Unregulated
Contaminant
Monitoring
(
UCM)
program
since
1993
(
USEPA,
1992).
Monitoring
ceased
for
small
public
water
systems
(
PWSs)
under
a
direct
final
rule
published
January
8,
1999
(
USEPA,
1999a),
and
ended
for
large
PWSs
with
promulgation
of
the
new
Unregulated
Contaminant
Monitoring
Regulation
(
UCMR)
issued
September
17,
1999
(
USEPA,
1999b)
and
effective
January
1,
2001.
At
the
time
the
UCMR
lists
were
developed,
the
Agency
concluded
there
were
adequate
monitoring
data
for
a
preliminary
regulatory
determination.
This
obviated
the
need
for
continued
monitoring
under
the
new
UCMR
list.
4­
4
Aldrin/
Dieldrin
 
February
2003
Data
Sources,
Data
Quality,
and
Analytical
Methods
Currently,
there
is
no
complete
national
record
of
unregulated
or
regulated
contaminants
in
drinking
water
from
PWSs
collected
under
SDWA.
Many
states
have
submitted
unregulated
contaminant
PWS
monitoring
data
to
EPA
databases,
but
there
are
issues
of
data
quality,
completeness,
and
representativeness.
Nonetheless,
a
significant
amount
of
state
data
are
available
for
UCM
contaminants
that
can
provide
estimates
of
national
occurrence.

The
National
Contaminant
Occurrence
Database
(
NCOD)
is
an
interface
to
the
actual
occurrence
data
stored
in
the
Safe
Drinking
Water
Information
System
(
Federal
version;
SDWIS/
FED)
and
can
be
queried
to
provide
a
summary
of
the
data
in
SDWIS/
FED
for
a
particular
contaminant.
The
drinking
water
occurrence
data
for
aldrin
presented
here
were
derived
from
monitoring
data
available
in
the
SDWIS/
FED
database.

The
data
in
this
report
have
been
reviewed,
edited,
and
filtered
to
meet
various
data
quality
objectives
for
the
purposes
of
this
analysis.
Hence,
not
all
data
from
a
particular
source
were
used,
only
data
meeting
the
quality
objectives
described
below
were
included.
The
sources
of
these
data,
their
quality
and
national
aggregation,
and
the
analytical
methods
used
to
estimate
a
given
contaminant's
national
occurrence
(
from
these
data)
are
discussed
in
this
section
(
for
further
details
see
USEPA,
2001a,
b).

UCM
Rounds
1
and
2
The
1987
UCM
contaminants
include
34
volatile
organic
compounds
(
VOCs)
(
USEPA,
1987).
Aldrin,
a
synthetic
organic
compound
(
SOC),
was
not
among
these
contaminants.
The
UCM
(
1987)
contaminants
were
first
monitored
coincident
with
the
Phase
I
regulated
contaminants,
during
the
1988
to
1992
period.
This
period
is
often
referred
to
as
"
Round
1"
monitoring.
The
monitoring
data
collected
by
the
PWSs
were
reported
to
the
states
(
as
primacy
agents),
but
there
was
no
protocol
in
place
to
report
these
data
to
EPA.
These
data
from
Round
1
were
collected
by
EPA
from
many
states
over
time
and
put
into
a
database
called
the
Unregulated
Contaminant
Information
System,
or
URCIS.

The
1993
UCM
contaminants
include
13
SOCs
and
1
inorganic
contaminant
(
IOC)
(
USEPA,
1991).
Monitoring
for
the
UCM
(
1993)
contaminants
began
coincident
with
the
Phase
II/
V
regulated
contaminants
in
1993
through
1998.
This
is
often
referred
to
as
"
Round
2"
monitoring.
The
UCM
(
1987)
contaminants
were
also
included
in
the
Round
2
monitoring.
As
with
other
monitoring
data,
PWSs
reported
these
results
to
the
states.
EPA,
during
the
past
several
years,
has
requested
that
all
states
submit
these
historic
data
to
EPA
and
they
are
now
stored
in
the
SDWIS/
FED
database.

Monitoring
and
data
collection
for
aldrin,
a
UCM
(
1993)
contaminant,
began
in
Round
2.
Therefore,
the
following
discussion
regarding
data
quality
screening,
data
management,
and
analytical
methods
focuses
on
SDWIS/
FED.
Discussion
of
the
URCIS
database
is
included
where
relevant,
but
it
is
worth
noting
that
the
various
quality
screening,
data
management,
and
4­
5
Aldrin/
Dieldrin
 
February
2003
analytical
processes
were
nearly
identical
for
the
two
databases.
For
further
details
on
the
two
monitoring
periods
as
well
as
the
databases
see
USEPA
(
2001a,
b).

Developing
a
Nationally
Representative
Perspective
The
Round
2
data
contain
contaminant
occurrence
data
from
a
total
of
35
primacy
entities
(
including
34
states
and
data
for
some
tribal
systems).
However,
data
from
some
states
are
incomplete
and
biased.
Furthermore,
the
national
representativeness
of
the
data
is
problematic
because
the
data
were
not
collected
in
a
systematic
or
random
statistical
framework.
These
state
data
could
be
heavily
skewed
to
low­
occurrence
or
high­
occurrence
settings.
Hence,
the
state
data
were
evaluated
based
on
pollution­
potential
indicators
and
the
spatial/
hydrologic
diversity
of
the
nation.
This
evaluation
enabled
the
construction
of
a
cross­
section
from
the
available
state
data
sets
that
provides
a
reasonable
representation
of
national
occurrence.

A
national
cross­
section
from
these
state
Round
2
contaminant
databases
was
established
using
the
approach
developed
for
the
EPA
report
A
Review
of
Contaminant
Occurrence
in
Public
Water
Systems
(
USEPA,
1999c).
This
approach
was
developed
to
support
occurrence
analyses
for
EPA's
Chemical
Monitoring
Reform
(
CMR)
evaluation.
It
was
supported
by
peer
reviewers
and
stakeholders.
The
approach
cannot
provide
a
"
statistically
representative"
sample
because
the
original
monitoring
data
were
not
collected
or
reported
in
an
appropriate
fashion.
However,
the
resultant
"
national
cross­
section"
of
states
should
provide
a
clear
indication
of
the
central
tendency
of
the
national
data.
The
remainder
of
this
section
provides
a
summary
description
of
how
the
national
cross­
section
for
the
SDWIS/
FED
(
Round
2)
database
was
developed.
The
details
of
the
approach
are
presented
in
other
documents
(
USEPA,
2001a;
USEPA,
2001b);
readers
are
referred
to
these
for
more
specific
information.

Cross­
Section
Development
As
a
first
step
in
developing
the
cross­
section,
the
state
data
contained
in
the
SDWIS/
FED
database
(
that
contains
the
Round
2
monitoring
results)
were
evaluated
for
completeness
and
quality.
Some
state
data
in
SDWIS/
FED
were
unusable
for
a
variety
of
reasons.
Some
states
reported
only
detections,
or
their
data
had
incorrect
units.
Datasets
only
including
detections
are
obviously
biased.
Other
problems
included
substantially
incomplete
data
sets
without
all
PWSs
reporting
(
USEPA,
2001a
Sections
II
and
III).

The
balance
of
the
states
remaining
after
the
data
quality
screening
were
then
examined
to
establish
a
national
cross­
section.
This
step
was
based
on
evaluating
the
states'
pollution
potential
and
geographic
coverage
in
relation
to
all
states.
Pollution
potential
is
considered
to
ensure
a
selection
of
states
that
represent
the
range
of
likely
contaminant
occurrence
and
a
balance
with
regard
to
likely
high
and
low
occurrence.
Geographic
consideration
is
included
so
that
the
wide
range
of
climatic
and
hydrogeologic
conditions
across
the
United
States
are
represented,
again
balancing
the
varied
conditions
that
affect
transport
and
fate
of
contaminants,
as
well
as
conditions
that
affect
naturally
occurring
contaminants
(
USEPA,
2001b
Sections
III.
A.
and
III.
B.).
4­
6
Aldrin/
Dieldrin
 
February
2003
The
cross­
section
states
were
selected
to
represent
a
variety
of
pollution
potential
conditions.
Two
primary
pollution
potential
indicators
were
used.
The
first
factor
selected
indicates
pollution
potential
from
manufacturing/
population
density
and
serves
as
an
indicator
of
the
potential
for
VOC
contamination
within
a
state.
Agriculture
was
selected
as
the
second
pollution
potential
indicator
because
the
majority
of
SOCs
of
concern
are
pesticides
(
USEPA,
2001b
Section
III.
A.).
The
50
individual
states
were
ranked
from
highest
to
lowest
based
on
the
pollution
potential
indicator
data.
For
example,
the
state
with
the
highest
ranking
for
pollution
potential
from
manufacturing
received
a
ranking
of
1
for
this
factor
and
the
state
with
the
lowest
value
was
ranked
as
number
50.
States
were
ranked
for
their
agricultural
chemical
use
status
in
a
similar
fashion.

The
states'
pollution
potential
rankings
for
each
factor
were
subdivided
into
four
quartiles
(
from
highest
to
lowest
pollution
potential).
The
cross­
section
states
were
chosen
from
all
quartiles
for
both
pollution
potential
factors
to
ensure
representation,
for
example,
from
the
following:
states
with
high
agrichemical
pollution
potential
rankings
and
high
manufacturing
pollution
potential
rankings;
states
with
high
agrichemical
pollution
potential
rankings
and
low
manufacturing
pollution
potential
rankings;
states
with
low
agrichemical
pollution
potential
rankings
and
high
manufacturing
pollution
potential
rankings;
and
states
with
low
agrichemical
pollution
potential
rankings
and
low
manufacturing
pollution
potential
rankings
(
USEPA,
2001b
Section
III.
B.).
In
addition,
some
secondary
pollution
potential
indicators
were
considered
to
further
ensure
that
the
cross­
section
states
included
the
spectrum
of
pollution
potential
conditions
(
high
to
low).
The
cross­
section
was
then
reviewed
for
geographic
coverage
throughout
all
sectors
of
the
United
States.

The
data
quality
screening,
pollution
potential
rankings,
and
geographic
coverage
analysis
established
a
national
cross­
section
of
20
Round
2
(
SDWIS/
FED)
states.
The
cross­
section
states
provide
a
good
representation
of
the
nation's
varied
climatic
and
hydrogeologic
regimes
and
the
breadth
of
pollution
potential
for
the
contaminant
groups
(
Figure
4­
1).

Cross­
Section
Evaluation
To
evaluate
and
validate
the
method
for
creating
the
national
cross­
sections,
the
method
was
used
to
create
smaller
state
subsets
from
the
24­
state,
Round
1
(
URCIS)
cross­
section
and
aggregations.
Again,
states
were
chosen
to
achieve
a
balance
from
the
quartiles
describing
pollution
potential,
and
a
balanced
geographic
distribution,
to
incrementally
build
subset
crosssections
of
various
sizes.
For
example,
the
Round
1
cross­
section
was
tested
with
subsets
of
4,
8
(
the
first
4
state
subset
plus
4
more
states),
and
13
(
8
state
subset
plus
5)
states.
Two
additional
cross­
sections
were
included
in
the
analysis
for
comparison:
a
cross­
section
composed
of
16
biased
states
eliminated
from
the
24
state
cross­
section
for
data
quality
reasons
and
a
crosssection
composed
of
all
40
Round
1
states
(
USEPA,
2001b
Section
III.
B.
1).
4­
7
Aldrin/
Dieldrin
 
February
2003
Figure
4­
1.
Geographic
Distribution
of
Cross­
Section
States
for
Round
2
(
SDWIS/
FED)

Round
2
(
SDWIS/
FED)

Alaska
Arkansas
Colorado
Kentucky
Maine
Maryland
Massachusetts
Michigan
Minnesota
Missouri
New
Hampshire
New
Mexico
North
Carolina
North
Dakota
Ohio
Oklahoma
Oregon
Rhode
Island
Texas
Washington
These
Round
1
incremental
cross­
sections
were
then
used
to
evaluate
occurrence
for
an
array
of
both
high
and
low
occurrence
contaminants.
The
comparative
results
illustrate
several
points.
The
results
are
quite
stable
and
consistent
for
the
8,
13,
and
24
state
cross­
sections.
They
are
much
less
for
the
4
state,
16
state
(
biased),
and
40
state
(
all
Round
1
states)
cross­
sections.
The
4
state
cross­
section
is
apparently
too
small
to
provide
balance
both
geographically
and
with
pollution
potential,
a
finding
that
concurs
with
past
work
(
USEPA,
1999c).
The
CMR
analysis
suggested
that
a
minimum
of
six
to
seven
states
was
needed
to
provide
balance
both
geographically
and
with
pollution
potential.
The
CMR
report
used
eight
states
out
of
the
available
data
for
its
nationally
representative
cross­
section
(
USEPA,
1999c).
The
16
state
and
40
state
cross­
sections,
both
including
biased
states,
provided
occurrence
results
that
were
unstable
and
inconsistent
for
a
variety
of
reasons
associated
with
their
data
quality
problems
(
USEPA,
2001b
Section
III.
B.
1).

The
8,
13,
and
24
state
cross­
sections
provide
very
comparable
results,
are
consistent,
and
are
usable
as
national
cross­
sections
to
provide
estimates
of
contaminant
occurrence.
Including
greater
data
from
more
states
improves
the
national
representation
and
the
confidence
in
the
results,
as
long
as
the
states
are
balanced
related
to
pollution
potential
and
spatial
coverage.
The
20
state
cross­
section
provides
the
best,
nationally
representative
cross­
section
for
the
Round
2
data.

Data
Management
and
Analysis
The
cross­
section
analyses
focused
on
occurrence
at
the
water
system
level;
i.
e.,
the
summary
data
presented
discuss
the
percentage
of
public
water
systems
with
detections,
not
the
percentage
of
samples
with
detections.
By
normalizing
the
analytical
data
to
the
system
level,
skewness
inherent
in
the
sample
data
is
avoided.
System
level
analysis
was
used
since
a
PWS
with
a
known
contaminant
problem
usually
has
to
sample
more
frequently
than
a
PWS
that
has
4­
8
Aldrin/
Dieldrin
 
February
2003
never
detected
the
contaminant.
Obviously,
the
results
of
a
simple
computation
of
the
percentage
of
samples
with
detections
(
or
other
statistics)
can
be
skewed
by
the
more
frequent
sampling
results
reported
by
the
contaminated
site.
This
level
of
analysis
is
conservative.
For
example,
a
system
need
only
have
a
single
sample
with
an
analytical
result
greater
than
the
Minimum
Reporting
Limit
(
MRL),
i.
e.,
a
detection,
to
be
counted
as
a
system
with
a
result
"
greater
than
the
MRL."

Also,
the
data
used
in
the
analyses
were
limited
to
only
those
data
with
confirmed
water
source
and
sampling
type
information.
Only
standard
SDWA
compliance
samples
were
used;
"
special"
samples,
or
"
investigation"
samples
(
investigating
a
contaminant
problem
that
would
bias
results),
or
samples
of
unknown
type
were
not
used
in
the
analyses.
Various
quality
control
and
review
checks
were
made
of
the
results,
including
follow­
up
questions
to
the
states
providing
the
data.
Many
of
the
most
intractable
data
quality
problems
encountered
occurred
with
older
data.
These
problematic
data
were,
in
some
cases,
simply
eliminated
from
the
analysis.
For
example,
when
the
number
of
data
with
problems
were
insignificant
relative
to
the
total
number
of
observations
they
were
dropped
from
the
analysis
(
for
further
details
see
Cadmus,
2000).

As
indicated
above,
Massachusetts
is
included
in
the
20­
state,
Round
2
national
crosssection
Noteworthy
for
SOCs
like
aldrin,
however,
Massachusetts
SOC
data
were
problematic.
Massachusetts
reported
Round
2
sample
results
for
SOCs
from
only
56
PWSs,
while
reporting
VOC
results
from
over
400
different
PWSs.
Massachusetts
SOC
data
also
contained
an
atypically
high
percentage
of
systems
with
analytical
detections
when
compared
to
all
other
states.
Through
communications
with
Massachusetts
data
management
staff,
it
was
learned
that
the
state's
SOC
data
were
incomplete
and
that
the
SDWIS/
FED
record
for
Massachusetts
SOC
data
were
also
incomplete.
For
instance,
the
SDWIS/
FED
Round
2
data
for
Massachusetts
indicates
18%
of
systems
reported
detections
of
aldrin.
The
average
percent
of
systems
with
detections
for
all
other
states
was
0.2%.
In
contrast,
Massachusetts
data
characteristics
and
quantities
for
IOCs
and
VOCs
were
reasonable
and
comparable
with
other
states'
results.
Therefore,
Massachusetts
was
included
in
the
group
of
20
SDWIS/
FED
Round
2
cross­
section
states
with
usable
data
for
IOCs
and
VOCs,
but
its
aldrin
(
SOC)
data
were
omitted
from
the
Round
2
cross­
section
occurrence
analyses
and
summaries
presented
in
this
report.

Occurrence
Analysis
To
evaluate
national
contaminant
occurrence,
a
two­
stage
analytical
approach
has
been
developed.
The
first
stage
of
analysis
provides
a
straightforward,
conservative,
broad
evaluation
of
occurrence
of
the
CCL
preliminary
regulatory
determination
priority
contaminants
as
described
above.
These
descriptive
statistics
are
summarized
here.
Based
on
the
findings
of
the
Stage
1
Analysis,
EPA
will
determine
whether
more
intensive
statistical
evaluations,
the
Stage
2
Analysis,
may
be
warranted
to
generate
national
probability
estimates
of
contaminant
occurrence
and
exposure
for
priority
contaminants.
(
For
details
on
this
two­
stage
analytical
approach
see
Cadmus,
2000.)

The
summary
descriptive
statistics
presented
in
Table
4­
2
for
aldrin
are
a
result
of
the
Stage
1
analysis
and
include
data
from
Round
2
(
SDWIS/
FED,
1993
to
1997)
cross­
section
states
4­
9
Aldrin/
Dieldrin
 
February
2003
(
excluding
Massachusetts).
Included
are
the
total
number
of
samples,
the
percent
samples
with
detections,
the
99th
percentile
concentration
of
all
samples,
the
99th
percentile
concentration
of
samples
with
detections,
and
the
median
concentration
of
samples
with
detections.
The
percentages
of
PWSs
and
population
served
indicate
the
proportion
of
PWSs
whose
analytical
results
showed
a
detection(
s)
of
the
contaminant
(
simple
detection,
>
MRL)
at
any
time
during
the
monitoring
period;
or
a
detection(
s)
greater
than
half
the
Health
Reference
Level
(
HRL);
or
a
detection(
s)
greater
than
the
HRL.
The
HRL,
0.002
µ
g/
L,
is
a
preliminary
estimated
health
effect
level
used
for
this
analysis.

Aldrin
is
classified
by
EPA
as
a
linear
carcinogen
and
would,
if
regulated,
have
a
MCLG
of
zero.
The
value
used
as
the
HRL
when
for
the
occurrence
evaluation
was
the
concentration
equivalent
to
a
one­
in­
a­
million
risk
based
on
the
EPA
cancer
slope
factor.

The
99th
percentile
concentration
is
used
here
as
a
summary
statistic
to
indicate
the
upper
bound
of
occurrence
values
because
maximum
values
can
be
extreme
values
(
outliers)
that
sometimes
result
from
sampling
or
reporting
error.
The
99th
percentile
concentration
is
presented
for
both
the
samples
with
only
detections
and
all
of
the
samples
because
the
value
for
the
99th
percentile
concentration
of
all
samples
is
below
the
Minimum
Reporting
Level
(
MRL)
(
denoted
by
"<"
in
Table
4­
2).
For
the
same
reason,
summary
statistics
such
as
the
95th
percentile
concentration
of
all
samples
or
the
median
(
or
mean)
concentration
of
all
samples
are
omitted
because
these
also
are
all
"<"
values.
This
is
the
case
because
only
0.006%
of
all
samples
recorded
detections
of
aldrin
in
Round
2.

As
a
simplifying
assumption,
a
value
of
half
the
MRL
is
often
used
as
an
estimate
of
the
concentration
of
a
contaminant
in
samples/
systems
whose
results
are
less
than
the
MRL.
For
a
contaminant
with
relatively
low
occurrence,
such
as
aldrin
in
drinking
water
occurrence
databases,
the
median
or
mean
value
of
the
occurrence
using
this
assumption
would
be
half
of
the
MRL
(
0.5
*
MRL).
However,
for
these
occurrence
data
this
is
not
straightforward.
For
Round
2,
states
have
reported
a
wide
range
of
values
for
the
MRLs.
This
is
in
part
related
to
state
data
management
differences,
as
well
as
real
differences
in
analytical
methods,
laboratories,
and
other
factors.

The
situation
can
cause
confusion
when
examining
descriptive
statistics
for
occurrence.
For
example,
most
Round
2
states
reported
non­
detections
simply
as
zeros
resulting
in
a
modal
MRL
value
of
zero.
By
definition
the
MRL
cannot
be
zero.
This
is
an
artifact
of
state
data
management
systems.
Because
a
simple
meaningful
summary
statistic
is
not
available
to
describe
the
various
reported
MRLs,
and
to
avoid
confusion,
MRLs
are
not
reported
in
the
summary
table
(
Table
4­
2).

In
Table
4­
2,
national
occurrence
is
estimated
by
extrapolating
the
summary
statistics
for
the
20
state
cross­
section
(
excluding
Massachusetts)
to
national
numbers
for
systems,
and
population
served
by
systems,
from
the
Water
Industry
Baseline
Handbook,
Second
Edition
(
USEPA,
2000).
From
the
handbook,
the
total
number
of
community
water
systems
(
CWSs)
plus
non­
transient,
non­
community
water
systems
(
NTNCWSs)
is
65,030,
and
the
total
population
served
by
CWSs
plus
NTNCWSs
is
213,008,182
persons
(
see
Table
4­
2).
To
arrive
4­
10
Aldrin/
Dieldrin
 
February
2003
at
the
national
occurrence
estimate
for
a
particular
cross­
section,
the
national
estimate
for
PWSs
(
or
population
served
by
PWSs)
is
simply
multiplied
by
the
percentage
for
the
given
summary
statistic
(
i.
e.,
the
national
estimate
for
the
total
number
of
PWSs
with
detections
[
11]
is
the
product
of
the
percentage
of
PWSs
with
detections
[
0.016%]
and
the
national
estimate
for
the
total
number
of
PWSs
[
65,030]).

Included
in
Table
4­
2
in
addition
to
the
cross­
section
data
results
are
results
and
national
extrapolations
from
all
Round
2
reporting
states.
The
data
from
the
biased
states
are
included
because
of
aldrin's
very
low
occurrence
in
drinking
water
samples
in
all
states.
For
contaminants
with
very
low
occurrence,
such
as
aldrin
where
very
few
states
have
detections,
any
occurrence
becomes
more
important,
relatively.
For
such
contaminants,
the
cross­
section
process
can
easily
miss
a
state
with
occurrence
that
becomes
more
important.
This
is
the
case
with
aldrin.

Extrapolating
only
from
the
cross­
section
states,
aldrin's
very
low
occurrence
clearly
underestimates
national
occurrence.
For
example,
while
data
from
biased
states
like
Alabama
(
reporting
100%
detections
>
HRL,
>
½
HRL,
and
>
MRL;
see
Appendix
A)
exaggerate
occurrence
because
only
systems
with
detections
reported
results,
their
detections
are
real
and
need
to
be
accounted
for
because
extrapolations
from
the
cross­
section
states
do
not
predict
enough
detections
in
the
biased
states.
Therefore,
results
from
all
reporting
Round
2
states,
including
the
biased
states,
are
also
used
here
to
extrapolate
to
a
national
estimate.
Using
the
biased
states'
data
should
provide
conservative
estimates,
likely
overestimates,
of
national
occurrence
for
aldrin.

As
exemplified
by
the
cross­
section
extrapolations
for
aldrin
and
dieldrin,
national
extrapolations
of
these
Stage
1
analytical
results
can
be
problematic,
especially
for
contaminants
with
very
low
occurrence,
because
the
State
data
used
for
the
cross­
section
are
not
a
strict
statistical
sample.
For
this
reason,
the
nationally
extrapolated
estimates
of
occurrence
based
on
Stage
1
results
are
not
presented
in
the
CCL
Federal
Register
Notice.
The
presentation
in
the
Federal
Register
Notice
of
only
the
actual
results
of
the
cross­
section
analysis
maintains
a
straight­
forward
presentation,
and
the
integrity
of
the
data,
for
stakeholder
review.
The
nationally
extrapolated
Stage
1
occurrence
values
are
presented
here,
however,
to
provide
additional
perspective.
A
more
rigorous
statistical
modeling
effort,
the
Stage
2
analysis,
could
be
conducted
on
the
cross­
section
data
(
Cadmus,
2001).
The
Stage
2
results
would
be
more
statistically
robust
and
more
suitable
to
national
extrapolation.
This
approach
would
provide
a
probability
estimate
and
would
also
allow
for
better
quantification
of
estimation
error.

Additional
Drinking
Water
Data
from
the
Corn
Belt
To
augment
the
SDWA
drinking
water
data
analysis
described
above,
and
to
provide
additional
coverage
of
the
corn
belt
states
where
aldrin
use
as
an
agricultural
insecticide
was
historically
high,
independent
analyses
of
SDWA
drinking
water
data
from
the
states
of
Iowa,
Illinois,
and
Indiana
are
reviewed
below.
The
Iowa
analysis
examined
SDWA
compliance
monitoring
data
from
surface
and
ground
water
PWSs
for
the
years
1988
to
1995
(
Hallberg
et
al.,
1996).
Illinois
and
Indiana
compliance
monitoring
data
for
surface
and
ground
water
PWSs
were
evaluated
mostly
for
the
years
after
1993,
though
some
earlier
data
were
also
included
(
USEPA,
4­
11
Aldrin/
Dieldrin
 
February
2003
1999c).
The
raw
water
data
from
Illinois
were
collected
from
rural,
private
supply
wells
(
Goetsch
et
al.,
1992).
Data
sources,
data
quality,
and
analytical
methods
for
these
analyses
are
described
in
the
respective
reports;
they
were
all
treated
similarly
to
the
data
quality
reviews
for
this
analysis.

Results
Occurrence
Estimates
The
percentages
of
PWSs
with
detections
are
very
low
(
Table
4­
2).
The
cross­
section
shows
only
approximately
0.02%
of
PWSs
(
approximately
11
PWSs
nationally)
experienced
detections
at
any
concentration
level
(>
MRL,
>
½
HRL,
and
>
HRL),
affecting
about
0.02%
of
the
population
served
(
approximately
40,000
to
50,000
people
nationally)
(
see
also
Figure
4­
2).
All
of
the
detections
were
in
systems
using
ground
water.
The
percentage
of
PWSs
(
or
population
served)
in
a
given
source
category
(
i.
e.,
ground
water)
with
detections
>
MRL,
>
½
HRL,
or
>
HRL
is
the
same
because
the
estimated
HRL
is
so
low
that
it
is
lower
than
the
MRL.
Hence,
any
detection
reported
is
also
greater
than
the
HRL.
While
concentrations
are
low
 
for
the
detections
the
median
concentration
is
0.58
µ
g/
L,
and
the
99th
percentile
concentration
is
0.69
µ
g/
L
 
these
values
are
greater
than
the
HRL.

As
noted
above,
because
of
the
very
low
occurrence,
the
cross­
section
states
yield
an
underestimate.
Hence,
all
data
are
used,
even
the
biased
data,
to
present
a
conservative
upper
bound
estimate.
Conservative
estimates
of
aldrin
occurrence
using
all
of
the
Round
2
reporting
states
still
show
relatively
low
detection
frequencies
(
Table
4­
2).
Approximately
0.2%
of
PWSs
(
estimated
at
138
PWSs
nationally)
experienced
detections
at
any
concentration
level
(>
MRL,
>
½
HRL,
and
>
HRL),
affecting
about
0.5%
of
the
population
served
(
1,052,000
people
nationally).
The
proportion
of
surface
water
PWSs
with
detections
was
greater
than
ground
water
systems.
Again
the
percentages
of
PWSs
(
or
populations
served)
with
detections
>
MRL,
>
½
HRL,
or
>
HRL
are
the
same
because
of
the
low
HRL.
The
median
concentration
of
detections
is
0.18
µ
g/
L,
and
the
99th
percentile
concentration
is
4.4
µ
g/
L.

The
Round
2
reporting
states
and
the
Round
2
national
cross­
section
show
a
proportionate
balance
in
PWS
source
waters
compared
to
the
national
inventory.
Nationally,
91%
of
PWSs
use
ground
water
(
and
9%
surface
waters).
Round
2
reporting
states
and
the
Round
2
national
cross­
section
show
87%
use
ground
water
(
and
13%
surface
waters).
The
relative
populations
served
are
not
as
comparable.
Nationally,
about
40%
of
the
population
is
served
by
PWSs
using
ground
water
(
and
60%
by
surface
water).
For
the
Round
2
cross­
section,
29%
of
the
cross­
section
population
is
served
by
ground
water
PWSs
(
and
71%
by
surface
water).
For
all
Round
2
reporting
states,
31%
of
the
population
is
served
by
ground
water
PWSs
(
and
69%
by
surface
water).
The
resultant
national
extrapolations
are
not
additive
as
a
consequence
of
these
disproportions.

Drinking
water
data
from
the
corn
belt
states
of
Iowa,
Indiana,
and
Illinois
also
show
very
low
occurrence
of
aldrin.
There
were
no
detections
of
the
pesticide
in
the
Iowa
or
Indiana
SDWA
aldrin
as
well.
Only
0.3%
of
all
sampled
wells
had
detections
at
a
reporting
limit
of
0.004
µ
g/
L
(
Goetsch
et
al.,
1992).
4­
12
Aldrin/
Dieldrin
 
February
2003
Table
4­
2.
Summary
Occurrence
Statistics
for
Aldrin
Frequency
Factors
20
State
Cross­
Section1
All
Reporting
States2
National
System
&
Population
Numbers3
Total
Number
of
Samples
31,083
41,565
­­
Percent
of
Samples
with
Detections
0.006%
0.132%
­­
99th
Percentile
Concentration
(
all
samples)
<
(
Non­
detect)
<
(
Non­
detect)
­­
Health
Reference
Level
0.002
µ
g/
L
0.002
µ
g/
L
­­
Minimum
Reporting
Level
(
MRL)
Variable4
Variable4
­­
99th
Percentile
Concentration
of
Detections
0.69
µ
g/
L
4.40
µ
g/
L
­­
Median
Concentration
of
Detections
0.58
µ
g/
L
0.18
µ
g/
L
­­
Total
Number
of
PWSs
12,165
15,123
65,030
Number
of
GW
PWSs
10,540
13,195
59,440
Number
of
SW
PWSs
1,625
1,928
5,590
Total
Population
47,708,156
58,979,361
213,008,182
Population
of
GW
PWSs
14,043,051
18,279,343
85,681,696
Population
of
SW
PWSs
33,665,105
40,700,018
127,326,486
Occurrence
by
System
National
Extrapolation5
PWSs
with
detections
(>
MRL)
0.016%
0.212%
11
138
Range
of
Cross­
Section
States
0
­
0.23%
0
­
100%
N/
A
N/
A
GW
PWSs
with
detections
0.019%
0.167%
11
99
SW
PWSs
with
detections
0.000%
0.519%
0
29
PWSs
>
1/
2
Health
Reference
Level
(
HRL)
0.016%
0.212%
11
138
Range
of
Cross­
Section
States
0
­
0.23%
0
­
100%
N/
A
N/
A
GW
PWSs
>
1/
2
Health
Reference
Level
0.019%
0.167%
11
99
SW
PWSs
>
1/
2
Health
Reference
Level
0.000%
0.519%
0
29
PWSs
>
Health
Reference
Level
0.016%
0.212%
11
138
Range
of
Cross­
Section
States
0
­
0.23%
0
­
100%
N/
A
N/
A
GW
PWSs
>
Health
Reference
Level
0.019%
0.167%
11
99
SW
PWSs
>
Health
Reference
Level
0.000%
0.519%
0
29
Occurrence
by
Population
Served
PWS
Population
Served
with
detections
0.018%
0.494%
39,000
1,052,000
Range
of
Cross­
Section
States
0
­
0.35%
0
­
100%
N/
A
N/
A
GW
PWS
Population
with
detections
0.062%
0.414%
53,000
355,000
SW
PWS
Population
with
detections
0.000%
0.530%
0
674,000
PWS
Population
Served
>
1/
2
Health
Reference
Level
0.018%
0.494%
39,000
1,052,000
Range
of
Cross­
Section
States
0
­
0.35%
0
­
100%
N/
A
N/
A
GW
PWS
Population
>
1/
2
Health
Reference
Level
0.062%
0.414%
53,000
355,000
SW
PWS
Population
>
1/
2
Health
Reference
Level
0.000%
0.530%
0
674,000
PWS
Population
Served
>
Health
Reference
Level
0.018%
0.494%
39,000
1,052,000
Range
of
Cross­
Section
States
0
­
0.35%
0
­
100%
N/
A
N/
A
GW
PWS
Population
>
Health
Reference
Level
0.062%
0.414%
53,000
355,000
SW
PWS
Population
>
Health
Reference
Level
0.000%
0.530%
0
674,000
1.
Summary
Results
based
on
data
from
20­
State
Cross­
Section
(
minus
Massachusetts),
from
SDWIS/
FED,
UCM
(
1993)
Round
2.
2.
Summary
Results
based
on
data
from
all
reporting
states
from
SDWIS/
FED,
UCM
(
1993)
Round
2.
3.
Total
PWS
and
population
numbers
are
from
EPA
March
2000
Water
Industry
Baseline
Handbook.
4.
See
text
for
discussion.
5.
National
extrapolations
are
from
the
20­
State
data
using
the
Baseline
Handbook
system
and
population
numbers.
­
PWS
=
Public
Water
Systems;
GW
=
Ground
Water;
SW
=
Surface
Water;
MRL
=
Minimum
Reporting
Level
(
for
laboratory
analyses);
Health
Reference
Level
=
Health
Reference
Level,
an
estimated
health
effect
level
used
for
preliminary
assessment
for
this
review;
N/
A
=
Not
Applicable."
­
The
Health
Reference
Level
(
HRL)
used
for
aldrin
is
0.002
:
g/
L.
This
is
a
draft
value
for
working
review
only.
­
Total
Number
of
Samples
=
the
total
number
of
analytical
records
for
aldrin.
­
99th
Percentile
Concentration
=
the
concentration
value
of
the
99th
percentile
of
either
all
analytical
results
or
just
the
detections
(
in
:
g/
L).
­
Median
Concentration
of
Detections
=
the
median
analytical
value
of
all
the
detections
(
analytical
results
greater
than
the
MRL)
(
in
:
g/
L).
­
Total
Number
of
PWSs
=
the
total
number
of
public
water
systems
with
records
for
aldrin.
­
Total
Population
Served
=
the
total
population
served
by
public
water
systems
with
records
for
aldrin.
­
%
PWS
with
detections,
%
PWS
>
½
Health
Reference
Level,
%
PWS
>
Health
Reference
Level
=
percent
of
the
total
number
of
public
water
systems
with
at
least
one
analytical
result
that
exceeded
the
MRL,
½
Health
Reference
Level,
Health
Reference
Level,
respectively.
­
%
PWS
Population
Served
with
detections,
%
PWS
Population
Served
>
½
Health
Reference
Level,
%
PWS
Population
Served
>
Health
Reference
Level
=
percent
of
the
total
population
served
by
PWSs
with
at
least
one
analytical
result
exceeding
the
MRL,
½
Health
Reference
Level,
or
the
Health
Reference
Level,
respectively.
4­
13
Aldrin/
Dieldrin
 
February
2003
Regional
Patterns
Occurrence
results
are
displayed
graphically
by
state
in
Figures
4­
2
and
4­
3
to
assess
whether
any
distinct
regional
patterns
of
occurrence
are
present.
Thirty­
four
states
reported
Round
2
data
but
seven
of
those
states
have
no
data
for
aldrin
(
Figure
4­
2).
Another
22
states
did
not
detect
aldrin.
The
remaining
five
states
have
detected
aldrin
in
drinking
water
and
are
generally
located
either
in
the
southern
United
States
or
the
Northeast
(
Figure
4­
2).
In
contrast
to
the
summary
statistical
data
presented
in
the
previous
section,
this
simple
spatial
analysis
includes
the
biased
Massachusetts
data.

The
simple
spatial
analysis
presented
in
Figures
4­
2
and
4­
3
suggests
that
special
regional
analyses
are
not
warranted.
The
State
of
Alabama
does,
however,
stand
out
as
having
relatively
high
occurrence
for
reasons
that
are
unclear.
While
there
is
a
weak
geographic
clustering
of
drinking
water
detections
in
a
few
southern
and
northeastern
states
(
including
the
State
of
Massachusetts'
biased
data),
this
is
partly
the
result
of
so
few
states
with
any
detections.
Further,
use
and
environmental
release
information
described
in
Chapter
3
of
this
report
indicates
that
aldrin
detections
are
more
widespread
than
the
drinking
water
data
suggest.
Two
out
of
the
three
TRI
states
(
Arkansas
and
Michigan)
that
reported
releases
of
aldrin
into
the
environment
did
not
report
detections
of
the
chemical
in
PWS
sampling.
Furthermore,
aldrin's
widespread
presence
in
the
environment
is
evidenced
by
detections
of
the
compound
in
hazardous
waste
sites
in
at
least
31
states
(
at
NPL
sites),
as
well
as
detections
in
site
samples
in
at
least
40
states
(
listed
in
ATSDR's
HazDat
[
ATSDR,
2000]).

4.1.3
Conclusion
Aldrin
is
an
insecticide
that
was
discontinued
for
all
uses
in
1987.
It
combats
insects
by
contact
or
ingestion,
and
was
used
primarily
on
corn
and
citrus
products,
as
well
as
for
general
crops
and
timber
preservation.
In
addition,
aldrin
was
used
for
termite­
proofing
plywood,
building
boards,
and
the
plastic
and
rubber
coverings
of
electrical
and
telecommunication
cables
(
ATSDR,
1993).
In
1972,
USEPA
cancelled
all
uses
of
aldrin
except
subsurface
ground
insertion
for
termite
control,
dipping
of
non­
food
plant
roots
and
tops,
and
moth­
proofing
in
closed­
system
manufacturing
processes.
This
cancellation
decision
was
finalized
in
1974,
and
in
1987,
the
manufacturer
voluntarily
cancelled
all
uses
(
ATSDR,
1993).
4­
14
Aldrin/
Dieldrin
 
February
2003
Aldrin
Detects
in
All
Round
2
States
States
not
in
Round
2
No
data
for
Aldrin
States
with
No
Detections
(
No
PWSs
>
MRL)
States
with
Detections
(
Any
PWSs
>
MRL)
All
States
Figure
4­
2.
States
With
PWSs
With
Detections
of
Aldrin
for
All
States
With
Data
in
SDWIS/
FED
(
Round
2)
4­
15
Aldrin/
Dieldrin
 
February
2003
*
State
of
Massachusetts
is
an
outlier
with
17.86%
PWSs
>
MRL
Aldrin
Occurrence
in
Cross­
section
States
States
not
in
Cross­
Section
No
data
for
Aldrin
0.00%
PWSs
>
MRL
0.01
­
1.00%
PWSs
>
MRL
>
1.00%
PWSs
>
MRL
*

Aldrin
Occurrence
in
Cross­
section
States
States
not
in
Cross­
Section
No
data
for
Aldrin
0.00%
PWSs
>
HRL
0.01
­
1.00%
PWSs
>
HRL
>
1.00%
PWSs
>
HRL
Figure
4­
3.
Round
2
Cross­
Section
States
With
PWSs
With
Detections
of
Aldrin
(
Any
PWSs
With
Results
Greater
than
the
Minimum
Reporting
Level
[
MRL];
Above)
and
Concentrations
Greater
than
the
Health
Reference
Level
(
HRL;
Below)
4­
16
Aldrin/
Dieldrin
 
February
2003
Aldrin
has
been
detected
at
very
low
frequencies
and
concentrations
in
bed
sediments
sampled
during
the
first
round
of
the
USGS
NAWQA
studies
and
in
ground
water
in
Illinois.
It
has
also
been
found
at
ATSDR
HazDat
and
CERCLA
NPL
sites
across
the
country.
Furthermore,
releases
have
been
reported
through
the
Toxic
Release
Inventory
(
TRI).

Aldrin
has
also
been
detected
in
PWS
samples
collected
under
the
Safe
Drinking
Water
Act
(
SDWA).
Occurrence
estimates
are
very
low
with
only
0.006%
of
all
cross­
section
samples
showing
detections.
Significantly,
the
values
for
the
99th
percentile
and
median
concentrations
of
all
cross­
section
samples
are
less
than
the
Minimum
Reporting
Level
(
MRL).
For
Round
2
cross­
section
samples
with
detections,
the
median
concentration
is
0.58
µ
g/
L
and
the
99th
percentile
concentration
is
0.69
µ
g/
L.
Systems
with
detections
constitute
only
0.02%
of
Round
2
cross­
section
systems
(
an
estimate
of
11
systems
nationally).
National
estimates
for
the
population
served
by
PWSs
with
detections
are
also
very
low
(
40,000
to
50,000),
and
are
the
same
for
all
categories
(>
MRL,
>
½
HRL,
>
HRL).
These
estimates
constitute
less
than
0.02%
of
the
national
population.
Using
more
conservative
estimates
of
occurrence
from
all
states
reporting
SDWA
Round
2
monitoring
data,
including
states
with
biased
data,
0.2%
of
the
nations
PWSs
(
approximately
138
systems)
and
0.5%
of
the
PWS
population
served
(
1,052,000
people)
may
be
estimated
to
have
detections
>
MRL,
>
½
HRL,
and
>
HRL.

Additional
SDWA
compliance
data
from
the
corn
belt
states
of
Iowa,
Indiana,
and
Illinois
examined
through
independent
analyses
support
the
drinking
water
data
analyzed
in
this
report.
There
were
no
detections
in
either
surface
or
ground
water
PWSs
in
the
states
of
Iowa
and
Indiana.
Illinois
reported
detections
only
from
surface
water
PWSs
with
1.8%
of
surface
water
systems,
and
0.1%
of
samples,
showing
detections.
The
99th
percentile
concentration
of
all
samples
was
below
the
reporting
level
and
the
maximum
concentration
was
2.4
µ
g/
L.
Furthermore,
in
a
survey
of
Illinois
rural,
private
water
supply
wells
aldrin
and
dealdrin
were
detected
in
only
0.3%
of
all
sampled
wells.
4­
17
Aldrin/
Dieldrin
 
February
2003
References
ATSDR.
2000.
Agency
for
Toxic
Substances
and
Disease
Registry.
Hazardous
Substance
Release
and
Health
Effects
Database.
Available
on
the
Internet
at:
http://
www.
atsdr.
cdc.
gov/
hazdat.
htm.
Last
modified
August
19,
2000.

ATSDR.
1993.
Agency
for
Toxic
Substances
and
Disease
Registry.
Toxicological
Profile
for
Aldrin/
Dieldrin
(
Update).
Atlanta:
Agency
for
Toxic
Substances
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Disease
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184
pp.

Cadmus.
2000.
Methods
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Estimating
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Exposure
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Water
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EPA
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July
25,
2000.

Cadmus.
2001.
Occurrence
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methodology
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year
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October
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Goetsch,
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D.,
D.
P.
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1992.
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18
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February
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of
unregulated
contaminant
monitoring
requirements
for
small
public
water
systems;
Final
Rule
and
Proposed
Rule.
Fed.
Reg.
64(
5):
1494
­
1498.
January
8.

USEPA.
1999b.
U.
S.
Environmental
Protection
Agency.
Revisions
to
the
unregulated
contaminant
monitoring
regulation
for
public
water
systems;
Final
Rule.
Fed.
Reg.
64(
180):
50556­
50620.
September
17.

USEPA.
1999c.
U.
S.
Environmental
Protection
Agency.
A
Review
of
Contaminant
Occurrence
in
Public
Water
Systems.
EPA
Report/
816­
R­
99/
006.
Office
of
Water.
78
pp.

USEPA.
1992.
U.
S.
Environmental
Protection
Agency.
Drinking
Water;
National
Primary
Drinking
Water
Regulations
 
Synthetic
Organic
Chemicals
and
Inorganic
Chemicals;
National
Primary
Drinking
Water
Regulations
Implementation.
Fed.
Reg.
57(
138):
31776­
31849.
July
17.

USEPA.
1991.
U.
S.
Environmental
Protection
Agency.
National
Primary
Drinking
Water
Regulations
­
Synthetic
Organic
Chemicals
and
Inorganic
Chemicals;
Monitoring
for
Unregulated
Contaminants;
National
Primary
Drinking
Water
Regulations
Implementation;
National
Secondary
Drinking
Water
Regulations;
Final
Rule.
Fed.
Reg.
56(
20)
3526­
3597.

USEPA.
1987.
U.
S.
Environmental
Protection
Agency.
National
Primary
Drinking
Water
Regulations­
Synthetic
Organic
Chemicals;
Monitoring
for
Unregulated
Contaminants;
Final
Rule.
Fed.
Reg.
52(
130):
25720.
July
8.

USGS.
2000.
U.
S.
Geological
Survey.
Pesticides
in
Stream
Sediment
and
Aquatic
Biota.
USGS
Fact
Sheet
FS­
092­
00.
4
pp.

USGS.
1999a.
U.
S.
Geological
Survey.
The
Quality
of
Our
Nation's
Waters:
Nutrients
and
Pesticides.
U.
S.
Geological
Survey
Circular
1225.
Reston,
VA:
United
States
Geological
Survey.
82
pp.

USGS.
1999b.
U.
S.
Geological
Survey.
Pesticides
Analyzed
in
NAWQA
Samples:
Use,
Chemical
Analyses,
and
Water­
Quality
Criteria.
PROVISIONAL
DATA
­­
SUBJECT
TO
REVISION.
Available
on
the
Internet
at:
http://
www.
water.
wr.
usgs.
gov/
pnsp/
anstrat.
Last
modified
August
20,
1999.
4­
19
Aldrin/
Dieldrin
 
February
2003
4.2
Dieldrin
4.2.1
Ambient
Occurrence
To
understand
the
presence
of
a
chemical
in
the
environment,
an
examination
of
ambient
occurrence
is
useful.
In
a
drinking
water
context,
ambient
water
is
source
water
existing
in
surface
waters
and
aquifers
before
treatment.
The
most
comprehensive
and
nationally
representative
data
describing
ambient
water
quality
in
the
United
States
are
being
produced
through
the
United
States
Geological
Survey's
(
USGS)
National
Water
Quality
Assessment
(
NAWQA)
program.
(
NAWQA,
however,
is
a
relatively
young
program
and
complete
national
data
are
not
yet
available
from
their
entire
array
of
sites
across
the
nation.)

Data
Sources
and
Methods
The
USGS
instituted
the
NAWQA
program
in
1991
to
examine
water
quality
status
and
trends
in
the
United
States.
NAWQA
is
designed
and
implemented
in
such
a
manner
to
enable
consistency
and
comparison
between
representative
study
basins
located
around
the
country,
facilitating
interpretation
of
natural
and
anthropogenic
factors
affecting
water
quality
(
Leahy
and
Thompson,
1994).

The
NAWQA
program
consists
of
59
significant
watersheds
and
aquifers
referred
to
as
"
study
units."
The
study
units
represent
approximately
two­
thirds
of
the
overall
water
usage
in
the
United
States
and
a
similar
proportion
of
the
population
served
by
public
water
systems.
Approximately
one­
half
of
the
nation's
land
area
is
represented
(
Leahy
and
Thompson,
1994).

To
facilitate
management
and
make
the
program
cost­
effective,
approximately
one­
third
of
the
study
units
at
a
time
engage
in
intensive
assessment
for
a
period
of
3
to
5
years.
This
is
followed
by
a
period
of
less
intensive
research
and
monitoring
that
lasts
between
5
and
7
years.
This
way
all
59
study
units
rotate
through
intensive
assessment
over
a
10­
year
period
(
Leahy
and
Thompson,
1994).
The
first
round
of
intensive
monitoring
(
1991
to
1996)
targeted
20
watersheds.
This
first
group
was
more
heavily
slanted
toward
agricultural
basins.
A
national
synthesis
of
results
from
these
study
units
focusing
on
pesticides
and
nutrients
has
been
compiled
and
analyzed
(
Kolpin
et
al.,
1998;
Larson
et
al.,
1999;
USGS,
1999).

Dieldrin
is
an
analyte
for
both
surface
and
ground
water
NAWQA
studies.
Two
of
the
first
20
study
basins
analyzed
in
the
pesticide
and
nutrient
national
synthesis
reports,
the
Central
Nebraska
Basins
and
the
White
River
Basin
in
Indiana,
are
located
in
the
corn
belt
where
dieldrin
use
was
heavy
during
the
1960s.
The
method
detection
limit
(
MDL)
for
dieldrin
is
0.001
µ
g/
L
(
Kolpin
et
al.,
1998),
substantively
lower
than
most
drinking
water
monitoring.
Additional
information
on
analytical
methods
used
in
the
NAWQA
study
units,
including
method
detection
limits,
are
described
by
Gilliom
and
others
(
in
press).

Dieldrin
is
an
organochlorine
insecticide.
As
a
group,
organochlorines
are
hydrophobic
and
resist
degradation.
Hydrophobic
("
water
hating")
compounds
have
low
water
solubilities
and
strong
tendencies
to
sorb
to
organic
material
in
sediments
and
accumulate
in
the
tissue
of
aquatic
biota,
where
they
can
persist
for
long
periods
of
time
(
ATSDR,
1993;
USGS,
2000).
4­
20
Aldrin/
Dieldrin
 
February
2003
Organochlorines
may
be
present
in
bed
sediments
and
tissues
of
aquatic
systems
even
when
they
are
undetectable
in
the
water
column
using
conventional
methods
(
Nowell,
1999).

To
determine
their
presence
in
hydrologic
systems
of
the
United
States,
the
NAWQA
program
has
investigated
organochlorine
pesticide
detections
in
bed
sediments
and
biotic
tissue,
focusing
on
the
organochlorine
insecticides
that
were
used
heavily
in
the
past
(
Nowell,
1999).
In
addition
to
its
own
commercial
production
and
use,
dieldrin
is
a
degradation
product
of
aldrin,
one
of
the
top
three
insecticides
used
for
agriculture
in
the
1960s
and
widely
used
to
kill
termites
in
structures
until
the
mid
1980s.
Given
this
history,
dieldrin
was
investigated
in
this
study
(
Nowell,
1999;
USGS,
1999).
Sampling
was
conducted
at
591
sites
from
1992
to
1995
in
the
20
NAWQA
study
units
first
intensively
assessed.
Details
regarding
sampling
techniques
and
analytical
methods
are
described
by
Nowell
(
1999).

Data
are
also
available
for
dieldrin
occurrence
in
surface
water
in
the
Mississippi
River
and
six
major
tributaries
draining
corn
belt
states
(
Goolsby
and
Battaglin,
1993).
These
data
are
the
result
of
a
USGS
regional
water
quality
investigation
and
details
regarding
sampling
and
analytical
methods
are
described
in
the
report.

Results
NAWQA
National
Synthesis
Detection
frequencies
and
concentrations
of
dieldrin
in
ambient
surface
and
ground
water
are
low,
especially
in
ground
water,
which
is
the
case
for
insecticides
in
general
(
Table
4­
3)
(
Kolpin
et
al.,
1998;
Miller
and
Wilber,
1999).
However,
using
a
common
reporting
limit
of
0.01
µ
g/
L,
dieldrin
is
the
most
commonly
detected
insecticide
in
ground
water
in
these
USGS
studies.
This
possibly
reflects
the
historically
heavy
use
of
aldrin
and
dieldrin
and
clearly
indicates
dieldrin's
environmental
persistence
(
Kolpin
et
al.,
1998;
Miller,
2000).
Also,
though
relatively
immobile
in
water
when
compared
to
newer
pesticides,
dieldrin
is
one
of
the
most
mobile
of
the
older
organochlorine
pesticides
(
USGS,
1999).

Dieldrin
detection
frequencies
are
considerably
higher
in
shallow
ground
water
in
urban
areas
when
compared
to
shallow
ground
water
in
agricultural
areas
(
Table
4­
3),
a
likely
consequence
of
the
more
recent
use
of
aldrin
and
dieldrin
as
a
termiticide
and
industrial
mothproofing
agent
until
the
mid­
1980s.
Agricultural
uses
were
discontinued
in
the
1970s.
Major
aquifers,
generally
deep,
have
very
low
detection
frequencies
and
concentrations
of
dieldrin.
Hydrophobic
compounds
have
high
sorption
potential
and
are
not
very
mobile
in
ground
water,
making
their
occurrence
in
deep
aquifers
unlikely.

In
streams,
detection
frequencies
are
higher
compared
to
ground
water
(
Table
4­
3).
Dieldrin's
chemical
characteristics,
chiefly
its
hydrophobicity,
make
it
less
likely
to
be
transported
to
the
subsurface
with
ground
water
recharge.
Instead,
dieldrin
sorbs
easily
to
sediments
and
biotic
tissues
and
may
persist
in
surface
water
environments
for
many
years
after
applications
have
ceased.
Differences
in
detection
frequencies
and
concentrations
between
urban
and
agricultural
settings
are
less
pronounced
for
streams
than
for
ground
water,
but
frequencies
and
concentrations
are
greater
for
streams
in
agricultural
settings.
4­
21
Aldrin/
Dieldrin
 
February
2003
The
concentrations
and
detection
frequencies
of
dieldrin
in
bed
sediments
and
biotic
tissues
are
considerably
higher
than
water,
although
the
median
concentration
of
all
samples
is
still
below
the
MDL
(
Table
4­
4).
Occurrence
of
dieldrin
is
highest
in
whole
fish,
highlighting
the
potential
for
it
to
bioaccumulate
(
Kolpin
et
al.,
1998).
The
trend
of
higher
concentrations
and
detection
frequencies
in
urban
environments
is
again
apparent
when
examining
dieldrin
Table
4­
3.
Dieldrin
Detections
and
Concentrations
in
Streams
and
Ground
Water1
Detection
Frequency
(%
Samples
$
MDL2)
Concentration
Percentiles
(
All
Samples;
µ
g/
L)

%
$
0.001
µ
g/
L
%
$
0.01
µ
g/
L
Median
95th
Maximum
Streams
urban
3.67%
1.83%
nd3
nd
0.016
integrator
3.27%
1.63%
nd
nd
0.015
agricultural
6.90%
3.90%
nd
0.007
0.027
all
sites
4.64%
2.39%
nd
nd
0.19
Ground
water
shallow
urban
5.65%
3.32%
nd
0.005
0.068
shallow
agricultural
0.97%
0.65%
nd
nd
0.057
major
aquifers
0.43%
0.21%
nd
nd
0.03
all
sites
1.42%
0.93%
nd
nd
0.068
1
USGS,
1998.

2
MDL
for
dieldrin
in
water
studies:
0.001
µ
g/
L.

3
Not
detected
in
concentration
greater
than
MDL.
4­
22
Aldrin/
Dieldrin
 
February
2003
Table
4­
4.
Dieldrin
Detections
and
Concentrations
in
Sediments,
Whole
Fish,
and
Bivalves
(
All
Sites)
1
Detection
Frequency
(%
Samples
>
MDL2)
Concentration
Percentiles
(
All
Samples;
µ
g/
kg
Dry
Weight)

Medium
95th
Maximum
sediments
13.7%
nd3
2.7
18
whole
fish
28.6%
nd
31.9
260
bivalves
6.4%
nd
6.4
20
1
Nowell,
1999.

2
MDL
for
dieldrin
in
sediments:
1
µ
g/
kg;
dieldrin
in
whole
fish
and
bivalves:
5
µ
g/
kg.

3
Not
detected
in
concentration
greater
than
MDL.

occurrence
across
various
land
use
settings
for
sediments
and
biotic
tissues.
Urban
areas
have
the
highest
detections
and
concentrations.
Occurrence
in
agricultural
and
mixed
land
use
settings
is
lower
and
approximately
equivalent.
Forest
and
rangeland
show
very
low
occurrence.
The
occurrence
of
a
toxic
compound
in
stream
sediments
is
pertinent
to
drinking
water
concerns
because
some
desorption
of
the
compound
from
sediments
into
water
will
occur
through
equilibrium
reactions,
although
in
very
low
concentrations.

While
concentrations
in
water
are
generally
low,
a
risk­
specific
dose
(
RSD)
criteria
of
0.02
µ
g/
L,
a
concentration
associated
with
a
cancer
risk
level
of
1
in
100,000
people,
was
exceeded
at
least
at
1
site
in
both
surface
and
ground
water
(
Kolpin
et
al.,
1998;
Larson
et
al.,
1999;
USGS,
1998).

Water
Quality
Investigations
from
the
Corn
Belt
A
USGS
regional
water
quality
investigation
provides
additional
information
on
the
occurrence
of
dieldrin
in
the
corn
belt.
For
surface
water
sampling
from
April
1991
to
March
1992
from
the
Mississippi
River
and
six
tributaries
draining
the
corn
belt,
8%
of
all
samples
and
71%
of
sites
had
detections
greater
than
the
reporting
limit
of
0.02
µ
g/
L.
The
maximum
concentration
was
approximately
0.03
µ
g/
L
(
Goolsby
and
Battaglin,
1993).

4.2.2
Drinking
Water
Occurrence
The
Safe
Drinking
Water
Act
(
SDWA),
as
amended
in
1986,
required
Public
Water
Systems
(
PWSs)
to
monitor
for
specified
"
unregulated"
contaminants,
conduct
monitoring
on
a
5­
year
cycle,
and
report
the
monitoring
results
to
the
states.
Unregulated
contaminants
do
not
have
an
established
or
proposed
National
Primary
Drinking
Water
Regulation
(
NPDWR),
but
they
are
contaminants
that
were
formally
listed
and
required
for
monitoring
under
federal
regulations.
The
intent
was
to
gather
scientific
information
on
the
occurrence
of
these
4­
23
Aldrin/
Dieldrin
 
February
2003
contaminants
to
enable
a
decision
as
to
whether
or
not
regulations
were
needed.
All
nonpurchased
community
water
systems
(
CWSs)
and
non­
purchased
non­
transient
non­
community
water
systems
(
NTNCWSs),
with
greater
than
150
service
connections,
were
required
to
conduct
this
unregulated
contaminant
monitoring.
Smaller
systems
were
not
required
to
conduct
this
monitoring
under
federal
regulations,
but
were
required
to
be
available
to
monitor
if
the
state
decided
such
monitoring
was
necessary.
Many
states
collected
data
from
smaller
systems.
Additional
contaminants
were
added
to
the
Unregulated
Contaminant
Monitoring
(
UCM)
program
in
1991
(
USEPA,
1991)
for
required
monitoring
that
began
in
1993
(
USEPA,
1992).

Dieldrin
has
been
monitored
under
the
SDWA
Unregulated
Contaminant
Monitoring
(
UCM)
program
since
1993
(
USEPA,
1992).
Monitoring
ceased
for
small
public
water
systems
(
PWSs)
under
a
direct
final
rule
published
January
8,
1999
(
USEPA,
1999a),
and
ended
for
large
PWSs
with
promulgation
of
the
new
Unregulated
Contaminant
Monitoring
Regulation
(
UCMR)
issued
September
17,
1999
(
USEPA,
1999b)
and
effective
January
1,
2001.
At
the
time
the
UCMR
lists
were
developed,
the
Agency
concluded
there
were
adequate
monitoring
data
for
a
regulatory
determination.
This
obviated
the
need
for
continued
monitoring
under
the
new
UCMR
list.

Data
Sources,
Data
Quality,
and
Analytical
Methods
Currently,
there
is
no
complete
national
record
of
unregulated
or
regulated
contaminants
in
drinking
water
from
PWSs
collected
under
SDWA.
Many
states
have
submitted
unregulated
contaminant
PWS
monitoring
data
to
EPA
databases,
but
there
are
issues
of
data
quality,
completeness,
and
representativeness.
Nonetheless,
a
significant
amount
of
state
data
are
available
for
UCM
contaminants
that
can
provide
estimates
of
national
occurrence.

The
National
Contaminant
Occurrence
Database
(
NCOD)
is
an
interface
to
the
actual
occurrence
data
stored
in
the
Safe
Drinking
Water
Information
System
(
Federal
version;
SDWIS/
FED)
and
can
be
queried
to
provide
a
summary
of
the
data
in
SDWIS/
FED
for
a
particular
contaminant.
The
drinking
water
occurrence
data
for
dieldrin
presented
here
were
derived
from
monitoring
data
available
in
the
SDWIS/
FED
database.

The
data
in
this
report
have
been
reviewed,
edited,
and
filtered
to
meet
various
data
quality
objectives
for
the
purposes
of
this
analysis.
Hence,
not
all
data
from
a
particular
source
were
used,
only
data
meeting
the
quality
objectives
described
below
were
included.
The
sources
of
these
data,
their
quality
and
national
aggregation,
and
the
analytical
methods
used
to
estimate
a
given
contaminant's
national
occurrence
(
from
these
data)
are
discussed
in
this
section
(
for
further
details
see
USEPA
[
2001a,
b]).

UCM
Rounds
1
and
2
The
1987
UCM
contaminants
include
34
volatile
organic
compounds
(
VOCs)
(
USEPA,
1987).
Dieldrin,
a
synthetic
organic
compound
(
SOC),
was
not
among
these
contaminants.
The
UCM
(
1987)
contaminants
were
first
monitored
coincident
with
the
Phase
I
regulated
contaminants,
during
the
1988
to
1992
period.
This
period
is
often
referred
to
as
"
Round
1"
monitoring.
The
monitoring
data
collected
by
the
PWSs
were
reported
to
the
states
(
as
primacy
4­
24
Aldrin/
Dieldrin
 
February
2003
agents),
but
there
was
no
protocol
in
place
to
report
these
data
to
EPA.
These
data
from
Round
1
were
collected
by
EPA
from
many
states
over
time
and
put
into
a
database
called
the
Unregulated
Contaminant
Information
System,
or
URCIS.

The
1993
UCM
contaminants
include
13
SOCs
and
1
inorganic
contaminant
(
IOC)
(
USEPA,
1991).
Monitoring
for
the
UCM
(
1993)
contaminants
began
coincident
with
the
Phase
II/
V
regulated
contaminants
in
1993
through
1998.
This
is
often
referred
to
as
"
Round
2"
monitoring.
The
UCM
(
1987)
contaminants
were
also
included
in
the
Round
2
monitoring.
As
with
other
monitoring
data,
PWSs
reported
these
results
to
the
states.
EPA,
during
the
past
several
years,
requested
that
the
states
submit
these
historic
data
to
EPA,
and
they
are
now
stored
in
the
SDWIS/
FED
database.

Monitoring
and
data
collection
for
dieldrin,
a
UCM
(
1993)
contaminant,
began
in
Round
2.
Therefore,
the
following
discussion
regarding
data
quality
screening,
data
management,
and
analytical
methods
focuses
on
SDWIS/
FED.
Discussion
of
the
URCIS
database
is
included
where
relevant,
but
it
is
worth
noting
that
the
various
quality
screening,
data
management,
and
analytical
processes
were
nearly
identical
for
the
two
databases.
For
further
details
on
the
two
monitoring
periods
as
well
as
the
databases
see
USEPA
(
2000a,
b).

Developing
a
Nationally
Representative
Perspective
The
Round
2
data
contain
contaminant
occurrence
data
from
a
total
of
35
primacy
entities
(
including
34
states
and
data
for
some
tribal
systems).
However,
data
from
some
states
are
incomplete
and
biased.
Furthermore,
the
national
representativeness
of
the
data
is
problematic
because
the
data
were
not
collected
in
a
systematic
or
random
statistical
framework.
These
state
data
could
be
heavily
skewed
to
low­
occurrence
or
high­
occurrence
settings.
Hence,
the
state
data
were
evaluated
based
on
pollution­
potential
indicators
and
the
spatial/
hydrologic
diversity
of
the
nation.
This
evaluation
enabled
the
construction
of
a
cross­
section
from
the
available
state
data
sets
that
provides
a
reasonable
representation
of
national
occurrence.

A
national
cross­
section
from
these
state
Round
2
contaminant
databases
was
established
using
the
approach
developed
for
the
EPA
report
A
Review
of
Contaminant
Occurrence
in
Public
Water
Systems
(
USEPA,
1999c).
This
approach
was
developed
to
support
occurrence
analyses
for
EPA's
Chemical
Monitoring
Reform
(
CMR)
evaluation.
It
was
supported
by
peer
reviewers
and
stakeholders.
The
approach
cannot
provide
a
"
statistically
representative"
sample
because
the
original
monitoring
data
were
not
collected
or
reported
in
an
appropriate
fashion.
However,
the
resultant
"
national
cross­
section"
of
states
should
provide
a
clear
indication
of
the
central
tendency
of
the
national
data.
The
remainder
of
this
section
provides
a
summary
description
of
how
the
national
cross­
section
for
the
SDWIS/
FED
(
Round
2)
database
was
developed.
The
details
of
the
approach
are
presented
in
other
documents
(
USEPA,
2001a;
USEPA
2001b).

Cross­
Section
Development
As
a
first
step
in
developing
the
cross­
section,
the
state
data
contained
in
the
SDWIS/
FED
database
(
that
contains
the
Round
2
monitoring
results)
were
evaluated
for
completeness
and
quality.
Some
state
data
in
SDWIS/
FED
were
unusable
for
a
variety
of
4­
25
Aldrin/
Dieldrin
 
February
2003
reasons.
Some
states
reported
only
detections,
or
their
data
had
incorrect
units.
Datasets
only
including
detections
are
obviously
biased.
Other
problems
included
substantially
incomplete
data
sets
without
all
PWSs
reporting
(
USEPA,
2001a
Sections
II
and
III).

The
balance
of
the
states
remaining
after
the
data
quality
screening
were
then
examined
to
establish
a
national
cross­
section.
This
step
was
based
on
evaluating
the
states'
pollution
potential
and
geographic
coverage
in
relation
to
all
states.
Pollution
potential
is
considered
to
ensure
a
selection
of
states
that
represent
the
range
of
likely
contaminant
occurrence
and
a
balance
with
regard
to
likely
high
and
low
occurrence.
Geographic
consideration
is
included
so
that
the
wide
range
of
climatic
and
hydrogeologic
conditions
across
the
United
States
are
represented,
again
balancing
the
varied
conditions
that
affect
transport
and
fate
of
contaminants,
as
well
as
conditions
that
affect
naturally
occurring
contaminants
(
USEPA,
2001b
Sections
III.
A.
and
III.
B.).

The
cross­
section
states
were
selected
to
represent
a
variety
of
pollution
potential
conditions.
Two
primary
pollution
potential
indicators
were
used.
The
first
factor
selected
indicates
pollution
potential
from
manufacturing/
population
density
and
serves
as
an
indicator
of
the
potential
for
VOC
contamination
within
a
state.
Agriculture
was
selected
as
the
second
pollution
potential
indicator
because
the
majority
of
SOCs
of
concern
are
pesticides
(
USEPA,
2001b
Section
III.
A.).
The
50
individual
states
were
ranked
from
highest
to
lowest
based
on
the
pollution
potential
indicator
data.
For
example,
the
state
with
the
highest
ranking
for
pollution
potential
from
manufacturing
received
a
ranking
of
1
for
this
factor
and
the
state
with
the
lowest
value
was
ranked
as
number
50.
States
were
ranked
for
their
agricultural
chemical
use
status
in
a
similar
fashion.

The
states'
pollution
potential
rankings
for
each
factor
were
subdivided
into
four
quartiles
(
from
highest
to
lowest
pollution
potential).
The
cross­
section
states
were
chosen
from
all
quartiles
for
both
pollution
potential
factors
to
ensure
representation,
for
example,
from
the
following:
states
with
high
agrichemical
pollution
potential
rankings
and
high
manufacturing
pollution
potential
rankings;
states
with
high
agrichemical
pollution
potential
rankings
and
low
manufacturing
pollution
potential
rankings;
states
with
low
agrichemical
pollution
potential
rankings
and
high
manufacturing
pollution
potential
rankings;
and
states
with
low
agrichemical
pollution
potential
rankings
and
low
manufacturing
pollution
potential
rankings
(
USEPA,
2001b
Section
III.
B.).
In
addition,
some
secondary
pollution
potential
indicators
were
considered
to
further
ensure
that
the
cross­
section
states
included
the
spectrum
of
pollution
potential
conditions
(
high
to
low).
The
cross­
section
was
then
reviewed
for
geographic
coverage
throughout
all
sectors
of
the
United
States.

The
data
quality
screening,
pollution
potential
rankings,
and
geographic
coverage
analysis
established
a
national
cross­
section
of
20
Round
2
(
SDWIS/
FED)
states.
The
cross­
section
states
provide
good
representation
of
the
nation's
varied
climatic
and
hydrogeologic
regimes
and
the
breadth
of
pollution
potential
for
the
contaminant
groups
(
Figure
4­
4).
4­
26
Aldrin/
Dieldrin
 
February
2003
Cross­
Section
Evaluation
To
evaluate
and
validate
the
method
for
creating
the
national
cross­
sections,
the
method
was
used
to
create
smaller
state
subsets
from
the
24­
state,
Round
1
(
URCIS)
cross­
section
and
aggregations.
Again,
states
were
chosen
to
achieve
a
balance
from
the
quartiles
describing
pollution
potential,
and
a
balanced
geographic
distribution,
to
incrementally
build
subset
crosssections
of
various
sizes.
For
example,
the
Round
1
cross­
section
was
tested
with
subsets
of
4,
8
(
the
first
4
state
subset
plus
4
more
states),
and
13
(
8
state
subset
plus
5)
states.
Two
additional
cross­
sections
were
included
in
the
analysis
for
comparison:
a
cross­
section
composed
of
16
biased
states
eliminated
from
the
24
state
cross­
section
for
data
quality
reasons
and
a
crosssection
composed
of
all
40
Round
1
states
(
USEPA,
2001b
Section
III.
B.
1).

These
Round
1
incremental
cross­
sections
were
then
used
to
evaluate
occurrence
for
an
array
of
both
high
and
low
occurrence
contaminants.
The
comparative
results
illustrate
several
points.
The
results
are
quite
stable
and
consistent
for
the
8,
13,
and
24
state
cross­
sections.
They
are
much
less
so
for
the
4
state,
16
state
(
biased),
and
40
state
(
all
Round
1
states)
cross­
sections.
The
4
state
cross­
section
is
apparently
too
small
to
provide
balance
both
geographically
and
Figure
4­
4.
Geographic
Distribution
of
Cross­
Section
States
for
Round
2
(
SDWIS/
FED)

Round
2
(
SDWIS/
FED)

Alaska
Arkansas
Colorado
Kentucky
Maine
Maryland
Massachusetts
Michigan
Minnesota
Missouri
New
Hampshire
New
Mexico
North
Carolina
North
Dakota
Ohio
Oklahoma
Oregon
Rhode
Island
Texas
Washington
4­
27
Aldrin/
Dieldrin
 
February
2003
with
pollution
potential,
a
finding
that
concurs
with
past
work
(
USEPA,
1999c).
The
CMR
analysis
suggested
that
a
minimum
of
6
to
7
states
were
needed
to
provide
balance
both
geographically
and
with
pollution
potential.
The
CMR
report
used
8
states
out
of
the
available
data
for
its
nationally
representative
cross­
section
(
USEPA,
1999c).
The
16
state
and
40
state
cross­
sections,
both
including
biased
states,
provided
occurrence
results
that
were
unstable
and
inconsistent
for
a
variety
of
reasons
associated
with
their
data
quality
problems
(
USEPA,
2001b
Section
III.
B.
1).

The
8,
13,
and
24
state
cross­
sections
provide
very
comparable
results,
are
consistent,
and
are
usable
as
national
cross­
sections
to
provide
estimates
of
contaminant
occurrence.
Including
greater
data
from
more
states
improves
the
national
representation
and
the
confidence
in
the
results,
as
long
as
the
states
are
balanced
related
to
pollution
potential
and
spatial
coverage.
The
20
state
cross­
section
provides
the
best,
nationally
representative
cross­
section
for
the
Round
2
data.

Data
Management
and
Analysis
The
cross­
section
analyses
focused
on
occurrence
at
the
water
system
level;
i.
e.,
the
summary
data
presented
discuss
the
percentage
of
public
water
systems
with
detections,
not
the
percentage
of
samples
with
detections.
By
normalizing
the
analytical
data
to
the
system
level,
skewness
inherent
in
the
sample
data
is
avoided.
System
level
analysis
was
used
since
a
PWS
with
a
known
contaminant
problem
usually
has
to
sample
more
frequently
than
a
PWS
that
has
never
detected
the
contaminant.
Obviously,
the
results
of
a
simple
computation
of
the
percentage
of
samples
with
detections
(
or
other
statistics)
can
be
skewed
by
the
more
frequent
sampling
results
reported
by
the
contaminated
site.
This
level
of
analysis
is
conservative.
For
example,
a
system
need
only
have
a
single
sample
with
an
analytical
result
greater
than
the
Minimum
Reporting
Limit
(
MRL),
i.
e.,
a
detection,
to
be
counted
as
a
system
with
a
result
"
greater
than
the
MRL."

Also,
the
data
used
in
the
analyses
were
limited
to
only
those
data
with
confirmed
water
source
and
sampling
type
information.
Only
standard
SDWA
compliance
samples
were
used;
"
special"
samples,
or
"
investigation"
samples
(
investigating
a
contaminant
problem
that
would
bias
results),
or
samples
of
unknown
type
were
not
used
in
the
analyses.
Various
quality
control
and
review
checks
were
made
of
the
results,
including
follow­
up
questions
to
the
states
providing
the
data.
Many
of
the
most
intractable
data
quality
problems
encountered
occurred
with
older
data.
These
problematic
data
were,
in
some
cases,
simply
eliminated
from
the
analysis.
For
example,
when
the
number
of
data
with
problems
were
insignificant
relative
to
the
total
number
of
observations
they
were
dropped
from
the
analysis
(
for
further
details
see
Cadmus
[
2000]).

As
indicated
above,
Massachusetts
is
included
in
the
20­
state,
Round
2
national
crosssection
(
Figure
4­
4).
However,
problematic
Massachusetts
data
for
SOCs
like
dieldrin
is
noteworthy.
Massachusetts
reported
Round
2
sample
results
for
SOCs
from
only
56
PWSs,
while
VOC
results
were
reported
from
over
400
different
PWSs.
Massachusetts
SOC
data
also
contained
an
atypically
high
percentage
of
systems
with
analytical
detections
when
compared
to
all
other
states.
Through
communications
with
Massachusetts
data
management
staff,
it
was
learned
that
the
state's
SOC
data
and
the
SDWIS/
FED
record
for
Massachusetts
SOC
data
were
4­
28
Aldrin/
Dieldrin
 
February
2003
incomplete.
For
instance,
the
SDWIS/
FED
Round
2
data
for
Massachusetts
indicates
18%
of
systems
reported
detections
of
dieldrin
while
the
average
for
all
other
states
was
0.4%.
In
contrast,
Massachusetts
data
characteristics
and
quantities
for
IOCs
and
VOCs
were
reasonable
and
comparable
with
other
states'
results.
Therefore,
Massachusetts
was
included
in
the
group
of
20
SDWIS/
FED
Round
2
cross­
section
states
with
usable
data
for
IOCs
and
VOCs,
but
its
dieldrin
(
SOC)
data
were
omitted
from
Round
2
cross­
section
occurrence
analyses
and
summaries
presented
in
this
report.

Occurrence
Analysis
To
evaluate
national
contaminant
occurrence,
a
two­
stage
analytical
approach
has
been
developed.
The
first
stage
of
analysis
provides
a
straightforward,
conservative,
broad
evaluation
of
occurrence
of
the
CCL
regulatory
determination
priority
contaminants
as
described
above.
These
descriptive
statistics
are
summarized
here.
Based
on
the
findings
of
the
Stage
1
Analysis,
EPA
will
determine
whether
more
intensive
statistical
evaluations,
the
Stage
2
Analysis,
may
be
warranted
to
generate
national
probability
estimates
of
contaminant
occurrence
and
exposure
for
priority
contaminants.
(
For
details
on
this
two
stage
analytical
approach
see
Cadmus
[
2000].)

The
summary
descriptive
statistics
presented
in
Table
4­
5
for
dieldrin
are
a
result
of
the
Stage
1
analysis
and
include
data
from
Round
2
(
SDWIS/
FED,
1993
to
1997)
cross­
section
states
(
minus
Massachusetts).
Included
are
the
total
number
of
samples,
the
percent
samples
with
detections,
the
99th
percentile
concentration
of
all
samples,
the
99th
percentile
concentration
of
samples
with
detections,
and
the
median
concentration
of
samples
with
detections.
The
percentages
of
PWSs
and
population
served
indicate
the
proportion
of
PWSs
whose
analytical
results
showed
a
detection(
s)
of
the
contaminant
(
simple
detection,
>
MRL)
at
any
time
during
the
monitoring
period;
or
a
detection(
s)
greater
than
half
the
HRL;
or
a
detection(
s)
greater
than
the
HRL.
The
HRL,
0.002
µ
g/
L,
is
a
preliminary
estimated
health
effect
level
used
for
this
analysis.

Dieldrin
is
classified
by
EPA
as
a
linear
carcinogen
and
would,
if
regulated,
have
a
MCLG
of
zero.
The
value
used
as
the
HRL
when
for
the
occurrence
evaluation
was
the
concentration
equivalent
to
a
one­
in­
a­
million
risk
based
on
the
EPA
cancer
slope
factor.

The
99th
percentile
concentration
is
used
here
as
a
summary
statistic
to
indicate
the
upper
bound
of
occurrence
values
because
maximum
values
can
be
extreme
values
(
outliers)
that
sometimes
result
from
sampling
or
reporting
error.
The
99th
percentile
concentration
is
presented
for
both
the
samples
with
only
detections
and
all
of
the
samples
because
the
value
for
the
99th
percentile
concentration
of
all
samples
is
below
the
Minimum
Reporting
Level
(
MRL)
(
denoted
by
"<"
in
Table
4­
5).
For
the
same
reason,
summary
statistics
such
as
the
95th
percentile
concentration
of
all
samples
or
the
median
(
or
mean)
concentration
of
all
samples
are
omitted
because
these
also
are
all
"<"
values.
This
is
the
case
because
only
0.064%
of
all
samples
recorded
detections
of
dieldrin
in
Round
2.

As
a
simplifying
assumption,
a
value
of
half
the
MRL
is
often
used
as
an
estimate
of
the
concentration
of
a
contaminant
in
samples/
systems
whose
results
are
less
than
the
MRL.
With
a
relatively
low
occurrence
contaminant
such
as
dieldrin
in
drinking
water
occurrence
databases,
4­
29
Aldrin/
Dieldrin
 
February
2003
the
median
or
mean
value
of
occurrence
using
this
assumption
would
be
half
the
MRL
(
0.5
*
MRL).
However,
for
these
occurrence
data
this
is
not
straightforward.
For
Round
2,
states
have
reported
a
wide
range
of
values
for
the
MRLs.
This
is
in
part
related
to
state
data
management
differences,
as
well
as
real
differences
in
analytical
methods,
laboratories,
and
other
factors.

The
situation
can
cause
confusion
when
examining
descriptive
statistics
for
occurrence.
For
example,
most
Round
2
states
reported
non­
detections
simply
as
zeros
resulting
in
a
modal
MRL
value
of
zero.
By
definition
the
MRL
cannot
be
zero.
This
is
an
artifact
of
state
data
management
systems.
Because
a
simple
meaningful
summary
statistic
is
not
available
to
describe
the
various
reported
MRLs,
and
to
avoid
confusion,
MRLs
are
not
reported
in
the
summary
table
(
Table
4­
5).

In
Table
4­
5,
national
occurrence
is
estimated
by
extrapolating
the
summary
statistics
for
the
20
state
cross­
section
(
minus
Massachusetts)
to
national
numbers
for
systems,
and
population
served
by
systems,
from
the
Water
Industry
Baseline
Handbook,
Second
Edition
(
USEPA,
2000).
From
the
handbook,
the
total
number
of
community
water
systems
(
CWSs)
plus
non­
transient,
non­
community
water
systems
(
NTNCWSs)
is
65,030
and
the
total
population
served
by
CWSs
plus
NTNCWSs
is
213,008,182
persons
(
Table
4­
5).
To
arrive
at
the
national
occurrence
estimate
for
the
cross­
section,
the
national
estimate
for
PWSs
(
or
population
served
by
PWSs)
is
simply
multiplied
by
the
percentage
for
the
given
summary
statistic
(
i.
e.,
the
national
estimate
for
the
total
number
of
PWSs
with
detections,
61,
is
the
product
of
the
percentage
of
PWSs
with
detections,
0.093%,
and
the
national
estimate
for
the
total
number
of
PWSs,
65,030).

Included
in
Table
4­
5
in
addition
to
the
cross­
section
data
results
are
results
and
national
extrapolations
from
all
Round
2
reporting
states.
The
data
from
the
biased
states
are
included
because
for
contaminants
with
very
low
occurrence,
such
as
dieldrin
where
few
states
have
detections,
any
occurrence
becomes
more
important,
relatively.
For
such
contaminants,
the
cross­
section
process
can
easily
miss
a
state
with
occurrence
that
becomes
more
important.
This
is
the
case
with
dieldrin.

Extrapolating
only
from
the
cross­
section
states,
dieldrin's
very
low
occurrence
probably
underestimates
national
occurrence.
For
example,
while
data
from
biased
states
like
Alabama
(
reporting
100%
detections
>
HRL,
>
½
HRL,
and
>
MRL;
see
Appendix
B)
exaggerate
occurrence
because
only
systems
with
detections
reported
results,
their
detections
are
real
and
need
to
be
accounted
for
because
extrapolations
from
the
cross­
section
states
do
not
predict
enough
detections
in
the
biased
states.
Therefore,
results
from
all
reporting
Round
2
states,
including
the
biased
states,
are
also
used
here
to
extrapolate
a
national
estimate.
Using
the
biased
states'
data
should
provide
conservative
estimates,
likely
overestimates,
of
national
occurrence
for
dieldrin.

Additional
Drinking
Water
Data
from
the
Corn
Belt
To
augment
the
SDWA
drinking
water
data
analysis
described
above
and
to
provide
additional
coverage
of
the
corn
belt
states
where
dieldrin
use
as
an
agricultural
insecticide
was
historically
high,
independent
analyses
of
SDWA
drinking
water
data
from
the
states
of
Iowa,
4­
30
Aldrin/
Dieldrin
 
February
2003
Illinois,
and
Indiana
were
reviewed.
Raw
water
monitoring
data
are
also
included
from
Illinois
community
water
supply
wells.

The
Iowa
analysis
examined
SDWA
compliance
monitoring
data
from
surface
and
ground
water
PWSs
for
the
years
1988
to
1995
(
Hallberg
et
al.,
1996).
Illinois
and
Indiana
compliance
monitoring
data
for
surface
and
ground
water
PWSs
were
evaluated
mostly
for
the
years
after
1993,
though
some
earlier
data
were
also
included
(
USEPA,
1999c).
The
raw
water
data
from
Illinois
were
collected
from
rural,
private
supply
wells
(
Goetsch
et
al.,
1992).
Data
sources,
data
quality,
and
analytical
methods
for
these
analyses
are
described
in
the
respective
reports;
they
were
all
treated
similarly
to
the
data
quality
reviews
for
this
analysis.

Results
Occurrence
Estimates
The
percentages
of
PWSs
with
detections
are
very
low
(
Table
4­
5).
The
cross­
section
shows
approximately
0.1%
of
PWSs
(
about
61
PWSs
nationally)
experienced
detections
at
any
concentration
level
(>
MRL,
>
½
HRL,
and
>
HRL),
affecting
less
than
0.1%
of
the
population
served
(
150,000
people
nationally,
see
Figure
4­
5).
The
percentage
of
PWSs
(
or
population
served)
in
a
given
source
category
(
i.
e.,
ground
water)
with
detections
>
MRL,
>
½
HRL,
and
>
HRL
is
the
same
because
the
estimated
HRL
is
so
low
that
it
is
less
than
the
MRL.
Hence,
any
detection
reported
is
greater
than
the
HRL.
Detection
frequencies
are
marginally
higher
for
surface
water
systems
when
compared
to
ground
water
systems.
While
concentrations
are
also
low
 
for
samples
with
detections
the
median
concentration
is
0.16
µ
g/
L
and
the
99th
percentile
concentration
is
1.36
µ
g/
L
 
these
values
are
greater
than
the
HRL.

As
noted
above,
because
of
the
very
low
occurrence,
the
cross­
section
states
yield
an
underestimate.
Hence,
all
data
are
used,
even
the
biased
data,
to
present
a
conservative
upper
bound
estimate.
Conservative
estimates
of
dieldrin
occurrence
using
all
of
the
Round
2
reporting
states
still
show
relatively
low
detection
frequencies
(
Table
4­
5).
Approximately
0.2%
of
PWSs
(
estimated
at
137
PWSs
nationally)
experienced
detections
at
any
concentration
level
(>
MRL,
>
½
HRL,
and
>
HRL),
affecting
about
0.4%
of
the
population
served
(
793,000
people
nationally).
The
proportion
of
surface
water
PWSs
with
detections
was
greater
than
ground
water
systems.
Again
the
percentages
of
PWSs
(
or
populations
served)
with
detections
>
MRL,
>
½
HRL,
or
>
HRL
are
the
same
because
of
the
low
HRL.
The
median
concentration
of
detections
is
0.42
µ
g/
L
and
the
99th
percentile
concentration
is
4.4
µ
g/
L.

The
Round
2
reporting
states
and
the
Round
2
national
cross­
section
show
a
proportionate
balance
in
PWS
source
waters
compared
to
the
national
inventory.
Nationally,
91%
of
PWSs
use
ground
water
(
and
9%
surface
waters).
Round
2
reporting
states
and
the
Round
2
national
cross­
section
show
88%
use
ground
water
(
and
12%
surface
waters).
The
relative
populations
served
are
not
as
comparable.
Nationally,
about
40%
of
the
population
is
served
by
PWSs
using
ground
water
(
and
60%
by
surface
water).
For
the
Round
2
cross­
section,
30%
of
the
cross­
section
population
is
served
by
ground
water
PWSs
(
and
70%
by
surface
water).
For
all
Round
2
reporting
states,
32%
of
the
population
is
served
by
ground
water
PWSs
4­
31
Aldrin/
Dieldrin
 
February
2003
(
and
68%
by
surface
water).
The
resultant
national
extrapolations
are
not
additive
as
a
consequence
of
these
disproportions.

Drinking
water
data
from
the
corn
belt
states
of
Iowa,
Indiana,
and
Illinois
also
show
very
low
occurrence
of
dieldrin.
There
were
no
detections
of
the
pesticide
in
the
Iowa
SDWA
compliance
monitoring
data
for
surface
or
ground
water
PWSs
(
Hallberg
et
al.,
1996).
While
Illinois
and
Indiana
also
had
no
detections
of
the
compound
in
ground
water
PWSs,
it
was
detected
in
surface
water
PWSs
in
those
states
(
USEPA,
1999c).
Occurrence
was
low
in
both
states:
1.8%
of
surface
water
systems
(
0.1%
of
samples)
showed
detections
in
Illinois;
and
2.1%
of
surface
water
systems
(
0.3%
of
samples)
showed
detections
in
Indiana.
For
Illinois
and
Indiana
surface
water
PWSs,
the
99th
percentile
concentrations
of
all
samples
were
below
the
reporting
level
and
the
maximum
concentrations
were
0.1
µ
g/
L
and
0.04
µ
g/
L,
respectively
(
USEPA,
1999c).
Furthermore,
in
a
survey
of
Illinois
rural,
private
water
supply
wells
only
1.6%
of
all
sampled
wells
had
detections
of
dieldrin
(
Goetsch
et
al.,
1992).

Regional
Patterns
Occurrence
results
are
displayed
graphically
by
state
in
Figures
4­
5
and
4­
6
to
assess
whether
any
distinct
regional
patterns
of
occurrence
are
present.
Thirty­
four
states
reported
Round
2
data
but
seven
of
those
states
have
no
data
for
dieldrin
(
Figure
4­
5).
Another
19
states
did
not
detect
dieldrin.
The
remaining
eight
states
detected
dieldrin
in
drinking
water
and
are
generally
located
either
in
the
southern
United
States
or
the
Northeast
(
Figure
4­
5).
In
contrast
to
the
summary
statistical
data
presented
in
the
previous
section,
this
simple
spatial
analysis
includes
the
biased
Massachusetts
data.

The
simple
spatial
analysis
presented
in
Figures
4­
5
and
4­
6
suggests
that
special
regional
analyses
are
not
warranted.
Alabama
does,
however,
stand
out
as
having
relatively
high
occurrence
for
reasons
that
are
unclear.
While
there
is
a
weak
geographic
clustering
of
drinking
water
detections
in
a
few
southern
and
northeastern
states
(
including
Massachusetts'
biased
data),
this
is
partly
the
result
of
so
few
states
with
any
detections.
Further,
use
and
environmental
release
information
(
Section
3)
and
ambient
water
quality
data
(
Section
4.2.1.2)
indicate
that
dieldrin
detections
are
more
widespread
than
the
drinking
water
data
suggest.
Detections
of
the
compound
in
hazardous
waste
sites
in
at
least
38
states
(
at
NPL
sites),
site
samples
in
at
least
40
states
(
listed
in
ATSDR's
HazDat
[
ATSDR,
2000]),
and
water,
sediment,
and
biotic
tissue
quality
data
from
the
NAWQA
program
provide
evidence
for
nationwide
occurrence.
4­
32
Aldrin/
Dieldrin
 
February
2003
Table
4­
5.
Summary
Occurrence
Statistics
for
Dieldrin
Frequency
Factors
20
State
Cross­
Section1
All
Reporting
States2
National
System
&
Population
Numbers3
Total
Number
of
Samples
29,603
40,055
­­
Percent
of
Samples
with
Detections
0.064%
0.135%
­­
99th
Percentile
Concentration
(
all
samples)
<
(
Non­
detect)
<
(
Non­
detect)
­­
Health
Reference
Level
0.002
µ
g/
L
0.002
µ
g/
L
­­
Minimum
Reporting
Level
(
MRL)
Variable4
Variable4
­­
99th
Percentile
Concentration
of
Detections
1.36
µ
g/
L
4.40
µ
g/
L
­­
Median
Concentration
of
Detections
0.16
µ
g/
L
0.42
µ
g/
L
­­
Total
Number
of
PWSs
11,788
14,725
65,030
Number
of
GW
PWSs
10,329
12,968
59,440
Number
of
SW
PWSs
1,459
1,757
5,590
Total
Population
45,784,187
56,909,027
213,008,182
Population
of
GW
PWSs
13,831,864
18,044,000
85,681,696
Population
of
SW
PWSs
31,952,323
38,865,027
127,326,486
Occurrence
by
System
National
Extrapolation5
PWSs
with
detections
(>
MRL)
0.093%
0.211%
61
137
Range
of
Cross­
Section
States
0
­
0.97%
0
­
100%
N/
A
N/
A
GW
PWSs
with
detections
0.087%
0.177%
52
105
SW
PWSs
with
detections
0.137%
0.455%
8
25
PWSs
>
1/
2
Health
Reference
Level
(
HRL)
0.093%
0.211%
61
137
Range
of
Cross­
Section
States
0
­
0.97%
0
­
100%
N/
A
N/
A
GW
PWSs
>
1/
2
Health
Reference
Level
0.087%
0.177%
52
105
SW
PWSs
>
1/
2
Health
Reference
Level
0.137%
0.455%
8
25
PWSs
>
Health
Reference
Level
0.093%
0.211%
61
137
Range
of
Cross­
Section
States
0
­
0.97%
0
­
100%
N/
A
N/
A
GW
PWSs
>
Health
Reference
Level
0.087%
0.177%
52
105
SW
PWSs
>
Health
Reference
Level
0.137%
0.455%
8
25
Occurrence
by
Population
Served
PWS
Population
Served
with
detections
0.070%
0.372%
150,000
793,000
Range
of
Cross­
Section
States
0
­
2.00%
0
­
100%
N/
A
N/
A
GW
PWS
Population
with
detections
0.146%
0.371%
125,000
318,000
SW
PWS
Population
with
detections
0.038%
0.372%
48,000
474,000
PWS
Population
Served
>
1/
2
Health
Reference
Level
0.070%
0.372%
150,000
793,000
Range
of
Cross­
Section
States
0
­
2.00%
0
­
100%
N/
A
N/
A
GW
PWS
Population
>
1/
2
Health
Reference
Level
0.146%
0.371%
125,000
318,000
SW
PWS
Population
>
1/
2
Health
Reference
Level
0.038%
0.372%
48,000
474,000
PWS
Population
Served
>
Health
Reference
Level
0.070%
0.372%
150,000
793,000
Range
of
Cross­
Section
States
0
­
2.00%
0
­
100%
N/
A
N/
A
GW
PWS
Population
>
Health
Reference
Level
0.146%
0.371%
125,000
318,000
SW
PWS
Population
>
Health
Reference
Level
0.038%
0.372%
48,000
474,000
1.
Summary
Results
based
on
data
from
20­
State
Cross­
Section
(
minus
Massachusetts),
from
SDWIS/
FED,
UCM
(
1993)
Round
2.
2.
Summary
Results
based
on
data
from
all
reporting
states
from
SDWIS/
FED,
UCM
(
1993)
Round
2;
see
text
for
further
discussion.
3.
Total
PWS
and
population
numbers
are
from
EPA
March
2000
Water
Industry
Baseline
Handbook.
4.
See
text
for
discussion.
5.
National
extrapolations
are
from
the
20­
State
data
using
the
Baseline
Handbook
system
and
population
numbers.
­
"
PWS
=
Public
Water
Systems;
GW
=
Ground
Water;
SW
=
Surface
Water;
MRL
=
Minimum
Reporting
Level
(
for
laboratory
analyses);
Health
Reference
Level
=
Health
Reference
Level,
an
estimated
health
effect
level
used
for
preliminary
assessment
for
this
review;
N/
A
=
Not
Applicable."
­
The
Health
Reference
Level
(
HRL)
used
for
dieldrin
is
0.002
:
g/
L.
This
is
a
draft
value
for
working
review
only.
­
Total
Number
of
Samples
=
the
total
number
of
analytical
records
for
dieldrin.
­
99th
Percentile
Concentration
=
the
concentration
value
of
the
99th
percentile
of
either
all
analytical
results
or
just
the
detections
(
in
:
g/
L).
­
Median
Concentration
of
Detections
=
the
median
analytical
value
of
all
the
detections
(
analytical
results
greater
than
the
MRL)
(
in
:
g/
L).
­
Total
Number
of
PWSs
=
the
total
number
of
public
water
systems
with
records
for
dieldrin.
­
Total
Population
Served
=
the
total
population
served
by
public
water
systems
with
records
for
dieldrin.
­
%
PWS
with
detections,
%
PWS
>
½
Health
Reference
Level,
%
PWS
>
Health
Reference
Level
=
percent
of
the
total
number
of
public
water
systems
with
at
least
one
analytical
result
that
exceeded
the
MRL,
½
Health
Reference
Level,
Health
Reference
Level,
respectively.
­
%
PWS
Population
Served
with
detections,
%
PWS
Population
Served
>
½
Health
Reference
Level,
%
PWS
Population
Served
>
Health
Reference
Level
=
percent
of
the
total
population
served
by
PWSs
with
at
least
one
analytical
result
exceeding
the
MRL,
½
Health
Reference
Level,
or
the
Health
Reference
Level,
respectively.
4­
33
Aldrin/
Dieldrin
 
February
2003
Dieldrin
Detections
in
All
Round
2
States
States
not
in
Round
2
No
data
for
Dieldrin
States
with
No
Detections
(
No
PWSs
>
MRL)
States
with
Detections
(
Any
PWSs
>
MRL)
All
States
Figure
4­
5.
States
With
PWSs
With
Detections
of
Dieldrin
for
All
States
With
Data
in
SDWIS/
FED
(
Round
2)
4­
34
Aldrin/
Dieldrin
 
February
2003
*
State
of
Massachusetts
is
an
outlier
with
18.18%
PWSs
>
MRL
Dieldrin
Occurrence
in
Cross­
section
States
States
not
in
Cross­
Section
No
data
for
Dieldrin
0.00%
PWSs
>
MRL
0.01
­
1.00%
PWSs
>
MRL
>
1.00%
PWSs
>
MRL
*

Dieldrin
Occurrence
in
Cross­
section
States
States
not
in
Cross­
Section
No
data
for
Dieldrin
0.00%
PWSs
>
HRL
0.01
­
1.00%
PWSs
>
HRL
>
1.00%
PWSs
>
HRL
Figure
4­
6.
Round
2
Cross­
Section
States
With
PWSs
With
Detections
of
Dieldrin
(
Any
PWS
With
Results
Greater
than
the
Minimum
Reporting
Level
[
MRL];
Above)
and
Concentrations
Greater
than
the
Health
Reference
Level
(
HRL;
Below)
4­
35
Aldrin/
Dieldrin
 
February
2003
4.2.3
Conclusion
Dieldrin
is
an
insecticide
that
was
discontinued
for
all
uses
in
1987.
It
combats
insects
by
contact
or
ingestion,
and
was
used
primarily
on
corn
and
citrus
products,
as
well
as
for
general
crops
and
timber
preservation.
In
addition,
dieldrin
was
used
for
termite­
proofing
plywood,
building
boards,
and
the
plastic
and
rubber
coverings
of
electrical
and
telecommunication
cables
(
ATSDR,
1993).
In
1972,
USEPA
cancelled
all
uses
of
dieldrin
except
subsurface
ground
insertion
for
termite
control,
dipping
of
non­
food
plant
roots
and
tops,
and
moth­
proofing
in
closed­
system
manufacturing
processes.
This
cancellation
decision
was
finalized
in
1974
and
in
1987
the
manufacturer
voluntarily
cancelled
all
uses
(
ATSDR,
1993).
Dieldrin
is
also
produced
by
the
environmental
degradation
of
aldrin,
an
insecticide
with
similar
uses
and
regulatory
history.

Dieldrin
has
been
detected
at
low
frequencies
and
concentrations
in
ground
and
surface
water
sampled
during
the
first
round
of
the
USGS
NAWQA
studies,
and
at
similar
frequencies
and
concentrations
in
surface
waters
of
the
Mississippi
River
and
major
tributaries.
Its
occurrence
is
greater
in
stream
bed
sediments
and
biotic
tissue.
Dieldrin
has
also
been
found
at
ATSDR
HazDat
and
CERCLA
NPL
sites
across
the
country.

Dieldrin
has
been
detected
in
PWS
samples
collected
under
the
SDWA.
Occurrence
estimates
are
very
low
with
only
0.06%
of
all
samples
showing
detections.
Significantly,
the
values
for
the
99th
percentile
and
median
concentrations
of
all
samples
are
less
than
the
MRL.
For
Round
2
samples
with
detections,
the
median
concentration
is
0.16
µ
g/
L
and
the
99th
percentile
concentration
is
1.36
µ
g/
L.
Systems
with
detections
constitute
approximately
0.1%
of
Round
2
systems.
National
estimates
for
the
population
served
by
PWSs
with
detections
are
also
low
(
150,000),
and
are
the
same
for
all
categories
(>
MRL,
>
½
HRL,
>
HRL).
These
estimates
are
less
than
0.1%
of
the
national
population.
Using
more
conservative
estimates
of
occurrence
from
all
states
reporting
SDWA
Round
2
monitoring
data,
including
states
with
biased
data,
0.2%
of
the
nations
PWSs
(
approximately
137
systems)
and
0.4%
of
the
PWS
population
served
(
793,000
people)
may
be
estimated
to
have
detections
>
MRL,
>
½
HRL,
and
>
HRL.

Additional
SDWA
compliance
data
from
the
corn
belt
states
of
Iowa,
Indiana,
and
Illinois
examined
through
independent
analyses
support
the
drinking
water
data
analyzed
in
this
report.
There
were
no
detections
in
either
surface
or
ground
water
PWSs
in
the
state
of
Iowa.
Illinois
and
Indiana
reported
detections
only
from
surface
water
PWSs
with
1.8%
of
Illinois'
surface
water
systems
(
0.1%
of
samples)
and
2.1%
of
Indiana's
surface
water
systems
(
0.3%
of
samples)
showing
detections.
For
Illinois
and
Indiana
surface
water
PWSs,
the
99th
percentile
concentrations
of
all
samples
were
below
the
reporting
level
and
the
maximum
concentrations
were
0.1
µ
g/
L
and
0.04
µ
g/
L,
respectively
(
USEPA,
1999c).
Moreover,
in
a
survey
of
Illinois
rural,
private
water
supply
wells
dieldrin
was
detected
in
only
1.6%
of
all
sampled
wells.
4­
36
Aldrin/
Dieldrin
 
February
2003
References
ATSDR.
2000.
Agency
for
Toxic
Substances
and
Disease
Registry.
Hazardous
Substance
Release
and
Health
Effects
Database.
Available
on
the
Internet
at:
http://
www.
atsdr.
cdc.
gov/
hazdat.
htm.
Last
modified
August
19,
2000.

ATSDR.
1993.
Agency
for
Toxic
Substances
and
Disease
Registry.
Toxicological
Profile
for
Aldrin/
Dieldrin
(
Update).
Atlanta:
Agency
for
Toxic
Substances
and
Disease
Registry.
184
pp.

Cadmus.
2000.
Methods
for
Estimating
Contaminant
Occurrence
and
Exposure
in
Public
Drinking
Water
Systems
in
Support
of
CCL
Determinations.
Draft
report
submitted
to
EPA
for
review
July
25,
2000.

Cadmus.
2001.
Occurrence
estimation
methodology
and
occurrence
findings
report
for
six­
year
regulatory
review.
Draft
report
submitted
to
EPA
for
review
October
5,
2001.

Gilliom,
R.
J.,
D.
K.
Mueller,
and
L.
H.
Nowell.
In
press.
Methods
for
comparing
water­
quality
conditions
among
National
Water­
Quality
Assessment
Study
Units,
1992­
95.
U.
S.
Geological
Survey
Open­
File
Report
97­
589.

Goetsch,
W.
D.,
D.
P.
McKenna,
and
T.
J.
Bicki.
1992.
Statewide
Survey
for
Agricultural
Chemicals
in
Rural,
Private
Water­
Supply
Wells
in
Illinois.
Springfield,
IL:
Illinois
Department
of
Agriculture,
Bureau
of
Environmental
Programs.
4
pp.

Goolsby,
D.
A.
and
W.
A.
Battaglin.
1993.
Occurrence,
distribution
and
transport
of
agricultural
chemicals
in
surface
waters
of
the
Midwestern
United
States.
In
Goolsby,
D.
A.,
L.
L.
Boyer,
and
G.
E.
Mallard,
compilers.
Selected
Papers
on
Agricultural
Chemicals
in
Water
Resources
of
the
Midcontinental
United
States.
U.
S.
Geological
Survey
Open­
File
Report
94­
418.
pp.
1­
25.

Hallberg,
G.
R.,
D.
G.
Riley,
J.
R.
Kantamneni,
P.
J.
Weyer,
and
R.
D.
Kelley.
1996.
Assessment
of
Iowa
Safe
Drinking
Water
Act
Monitoring
Data:
1988­
1995.
Research
Report
No.
97­
1.
Iowa
City:
The
University
of
Iowa
Hygienic
Laboratory.
132
pp.

Kolpin,
D.
W.,
J.
E.
Barbash,
and
R.
J.
Gilliom.
1998.
Occurrence
of
pesticides
in
shallow
groundwater
of
the
United
States:
initial
results
from
the
National
Water
Quality
Assessment
Program.
Environ.
Sci.
Technol.
32:
558­
566.

Larson,
S.
J.,
R.
J.
Gilliom,
and
P.
D.
Capel.
1999.
Pesticides
in
Streams
of
the
United
States­­
Initial
Results
from
the
National
Water­
Quality
Assessment
Program.
U.
S.
Geological
Survey
Water­
Resources
Investigations
Report
98­
4222.
92
pp.
Available
on
the
Internet
at:
URL:
http://
water.
wr.
usgs.
gov/
pnsp/
rep/
wrir984222/.

Leahy,
P.
P.
and
T.
H.
Thompson.
1994.
The
National
Water­
Quality
Assessment
Program.
U.
S.
Geological
Survey
Open­
File
Report
94­
70.
4
pp.
Available
on
the
Internet
at:
http://
water.
usgs.
gov/
nawqa/
NAWQA.
OFR94­
70.
html
Last
updated
August
23,
2000.
4­
37
Aldrin/
Dieldrin
 
February
2003
Miller,
T.
2000.
Selected
Findings
and
Current
Perspectives
on
Urban
Water
Quality­
The
National
Water
Quality
Assessment
(
NAWQA)
Program
of
the
U.
S.
Geological
Survey.
Paper
presented
to
the
NAWQA
National
Liaison
Committee,
June
13,
2000.
8
pp.

Miller,
T.
L.
and
W.
G.
Wilber.
1999.
Emerging
Drinking
Water
Contaminants:
Overview
and
Role
of
the
National
Water
Quality
Assessment
Program
(
Ch
2.).
In:
Identifying
Future
Drinking
Water
Contaminants.
Washington,
D.
C.:
National
Academy
Press.

Nowell,
L.
1999.
National
Summary
of
Organochlorine
Detections
in
Bed
Sediment
and
Tissues
for
the
1991
NAWQA
Study
Units.
Available
on
the
Internet
at:
http://
water.
wr.
usgs.
gov/
pnsp/
rep/
bst/
Last
updated
October
18,
1999.

USEPA.
2001a.
Analysis
of
national
occurrence
of
the
1998
Contaminant
Candidate
List
regulatory
determination
priority
contaminants
in
public
water
systems.
Office
of
Water.
EPA
report
815­
D­
01­
002.
77
pp.

USEPA.
2001b.
Occurrence
of
unregulated
contaminants
in
public
water
systems:
An
initial
assessment.
Office
of
Water.
EPA
report
815­
P­
00­
001.
Office
of
Water.
50
pp.

USEPA.
2000.
U.
S.
Environmental
Protection
Agency.
Water
Industry
Baseline
Handbook,
Second
Edition
(
Draft).
March
17,
2000.

USEPA.
1999a.
U.
S.
Environmental
Protection
Agency.
Suspension
of
unregulated
contaminant
monitoring
requirements
for
small
public
water
systems;
Final
Rule
and
Proposed
Rule.
Fed.
Reg.
64(
5):
1494­
1498.
January
8.

USEPA.
1999b.
U.
S.
Environmental
Protection
Agency.
Revisions
to
the
unregulated
contaminant
monitoring
regulation
for
public
water
systems;
Final
Rule.
Fed.
Reg.
64(
180):
50556
­
50620.
September
17.

USEPA.
1999c.
U.
S.
Environmental
Protection
Agency.
A
Review
of
Contaminant
Occurrence
in
Public
Water
Systems.
EPA
Report/
816­
R­
99/
006.
Office
of
Water.
78
pp.

USEPA.
1992.
U.
S.
Environmental
Protection
Agency.
Drinking
Water;
National
Primary
Drinking
Water
Regulations
 
Synthetic
Organic
Chemicals
and
Inorganic
Chemicals;
National
Primary
Drinking
Water
Regulations
Implementation.
Fed.
Reg.
57(
138):
31776
­
31849.
July
17.

USEPA.
1991.
National
Primary
Drinking
Water
Regulations
­
Synthetic
Organic
Chemicals
and
Inorganic
Chemicals;
Monitoring
for
Unregulated
Contaminants;
National
Primary
Drinking
Water
Regulations
Implementation;
National
Secondary
Drinking
Water
Regulations;
Final
Rule.
Fed.
Reg.
56(
20)
3526­
3597.
January
30.

USEPA.
1987.
National
Primary
Drinking
Water
Regulations­
Synthetic
Organic
Chemicals;
Monitoring
for
Unregulated
Contaminants;
Final
Rule.
Fed.
Reg.
52(
130):
25720.
July
8.
4­
38
Aldrin/
Dieldrin
 
February
2003
USGS.
2000.
U.
S.
Geological
Survey.
Pesticides
in
Stream
Sediment
and
Aquatic
Biota.
USGS
Fact
Sheet
FS­
092­
00.
4
pp.

USGS.
1999.
U.
S.
Geological
Survey.
The
Quality
of
Our
Nation's
Waters:
Nutrients
and
Pesticides.
U.
S.
Geological
Survey
Circular
1225.
Reston,
VA:
United
States
Geological
Survey.
82
pp.

USGS.
1998.
U.
S.
Geological
Survey.
Pesticides
in
Surface
and
Ground
Water
of
the
United
States:
Summary
of
Results
of
the
National
Water
Quality
Assessment
Program
(
NAWQA).
PROVISIONAL
DATA
­­
SUBJECT
TO
REVISION.
Available
on
the
Internet
at:
http://
water.
wr.
usgs.
gov/
pnsp/
allsum/.
Last
modified
October
9,
1998.
5­
1
Aldrin/
Dieldrin
 
February
2003
5.0
EXPOSURE
FROM
ENVIRONMENTAL
MEDIA
OTHER
THAN
WATER
This
section
summarizes
human
population
exposures
to
aldrin
and
dieldrin
from
food,
air,
and
soil.
The
primary
purpose
is
to
estimate
average
daily
intakes
of
aldrin
and
dieldrin
by
members
of
the
general
public.
When
exposure
data
on
subpopulations
were
located,
such
as
occupationally
exposed
persons,
these
data
were
summarized
and
included
in
this
section.

5.1
Exposure
from
Food
Aldrin
and
dieldrin
have
been
used
for
pest
control
on
crops
such
as
corn,
and
citrus
products.
Aldrin
is
readily
converted
to
dieldrin,
which
is
persistent
in
the
environment.
Although
the
use
of
aldrin
and
dieldrin
on
crops
was
cancelled
in
1974,
soil
residues
from
past
uses
persist,
and
may
be
taken
up
by
crops.
Dieldrin
additionally
bioconcentrates
and
biomagnifies
through
terrestrial
and
aquatic
food
chains.
Thus,
the
general
population
may
be
exposed
to
aldrin
or
dieldrin
through
diet
(
ATSDR,
2000).

5.1.1
Exposures
of
the
General
Population
Concentrations
in
Non­
Fish
Food
Items
Aldrin
During
1981
through
1992,
the
U.
S.
Food
and
Drug
Administration
(
FDA)
conducted
a
Market
Basket
Study
to
evaluate
concentrations
of
pesticides
in
234
different
food
items.
Table
5­
1
summarizes
aldrin
concentrations
detected
in
these
foods.
Aldrin
was
detected
in
5
food
items
at
concentrations
ranging
from
0.0009
to
0.002
mg/
kg
food.
The
mean
concentration
for
all
positive
samples
was
0.0016
mg/
kg
(
KAN­
DO
Office
and
Pesticides
Team,
1995).

Agriculture
and
Agri­
Food
Canada
(
Neidert
and
Saschenbrecker,
1996)
analyzed
21,982
randomly
sampled
domestic
and
imported
food
and
vegetable
commodities
for
pesticide
residues
between
1992
and
1994.
Aldrin
was
not
detected
in
any
domestically
produced
fruits
or
vegetables,
but
was
detected
in
one
sample
of
imported
tomatoes
at
<
0.05
mg/
kg.
Aldrin
was
not
detected
in
any
food
items
during
the
1985
survey
(
Davies,
1988).

Kannan
et
al.
(
1994)
reviewed
data
on
aldrin
and
dieldrin
residues
in
food
in
South
and
Southeast
Asia
and
in
the
South
Pacific
Islands.
Aldrin
was
detected
in
several
food
items
collected
throughout
India
during
the
period
of
1975
through
1989.
Vegetables,
oils,
and
food
grains
contained
<
0.01
to
0.04
mg/
kg,
0.01
to
1.1
mg/
kg,
and
0.05
to
0.1
mg/
kg
aldrin,
respectively.

In
1990,
Kannan
et
al.
(
1994)
analyzed
food
items
collected
from
various
metropolitan
locations
in
Australia
for
organochlorine
pesticides.
The
highest
aldrin
concentrations
were
detected
in
pulses
and
dairy
products
at
levels
of
2.8
×
10­
3
and
8.9
×
10­
4
mg/
kg
wet
weight,
respectively.
Aldrin
was
also
detected
in
cereals
(
3
×
10­
5
mg/
kg),
oils
(
1.5
×
10­
4
mg/
kg),
vegetables
(
0.01
mg/
kg),
fruits
(<
0.01
mg/
kg),
and
meat
(
3.0
×
10­
4
mg/
kg).
5­
2
Aldrin/
Dieldrin
 
February
2003
Table
5­
1.
Aldrin
and
Dieldrin
in
Domestic
Food
Items
1981
to
19921
Type
of
Food
Mean
Dieldrin
Concentrations
(
mg/
kg
food)
and
Number
of
Positive
Samples
(
N)
Mean
Aldrin
Concentrations
(
mg/
kg
food)
and
Number
of
Positive
Samples
(
N)

Condiments,
Fats,
and
Sweetners
0.0011­
0.005
(
55)
­­

Dairy
0.0003­
0.0061
(
163)
­­

Desserts
0.0004­
0.0048
(
96)
0.0009
(
1)

Fruits
0.0005­
0.004
(
21)
­­

Grains
0.0003­
0.002
(
2)
0.002
(
1)

Infant
Food
(
strained
junior
foods
in
jars)
0.0003­
0.0051
(
36)
­­

Meat,
Poultry,
Fish
and
Eggs
0.0005­
0.002
(
195)
0.002
(
1)

Mixed
Foods
0.0006­
0.002
(
49)
­­

Soup
0.0004­
0.0008
(
9)
0.001
(
1)

Vegetables
and
Vegetable
Products
0.0002­
0.0108
(
210)
0.002
(
1)

1
Source:
KAN­
DO
Office
and
Pesticides
Team,
1995.

Milk
samples
collected
during
1990
through
1991
from
63
metropolitan
locations
throughout
the
United
States
did
not
contain
aldrin
residues
above
the
detection
limit
of
0.0005
ppm
(
Trotter
and
Dickerson,
1993).

During
FDA
Regulatory
Monitoring
1985­
1991
(
Yess
et
al.,
1993)
of
adult
foods
eaten
by
infants,
1
of
735
imported
orange
samples
analyzed
contained
trace
levels
of
aldrin.
However,
aldrin
was
not
detected
in
domestic
samples
of
adult
food
items
eaten
by
infants
analyzed
in
the
same
FDA
Regulatory
Monitoring
Survey
1985­
1991.
Infant
foods
analyzed
during
FDA
Total
Diet
Study
1985­
1991
(
Yess
et
al.,
1993)
and
Market
Basket
Survey
1981­
1991
(
KAN­
DO
Office
and
Pesticides
Team,
1995)
sampling
did
not
contain
detectable
levels
of
aldrin.
5­
3
Aldrin/
Dieldrin
 
February
2003
Dieldrin
Table
5­
1
summarizes
dieldrin
concentrations
in
various
food
items
analyzed
during
1981
through
1992
as
part
of
the
FDA's
Market
Basket
Study
(
KAN­
DO
Office
and
Pesticides
Team,
1995).
Dieldrin
was
detected
in
117
of
234
different
food
items
at
concentrations
ranging
from
0.0002
to
0.0087
mg/
kg.
The
mean
dieldrin
concentration
for
all
positive
samples
was
0.0015
mg/
kg.
The
highest
dieldrin
concentrations
were
detected
in
squash
(
0.0087
mg/
kg)
and
butter
(
0.0061
mg/
kg)
samples.
Cauliflower
(
0.0002
mg/
kg),
soup,
canned
beets,
and
red
beans
(
0.0004
mg/
kg)
had
the
lowest
dieldrin
concentrations.

In
1992
and
1994,
dieldrin
was
detected
in
both
domestic
and
imported
food
and
vegetable
commodities
analyzed
by
Agriculture
and
Agri­
Food
Canada
(
Neidert
and
Saschenbrecker,
1996).
Six
of
the
5,784
domestically
produced
fruits
and
vegetables
had
dieldrin
residues
ranging
from
<
0.05
to
0.10
mg/
kg.
Of
the
16,198
imported
fruits
and
vegetables
sampled,
7
had
dieldrin
levels
ranging
from
<
0.05
to
0.10
mg/
kg.
One
of
the
1,858
imported
oranges
contained
0.50
mg/
kg
dieldrin.
A
1985
Canadian
study
reported
higher
levels
of
dieldrin
residues
in
fruits
and
vegetables,
which
ranged
from
0.11
to
23.0
:
g/
kg
(
Davies,
1988).

Dieldrin
has
been
detected
in
various
meats.
Beef,
chicken,
lamb,
and
pork
samples
bought
from
butcher
shops
in
Australia
during
1990
contained
a
mean
dieldrin
concentration
of
5.1
×
10­
3
mg/
kg
wet
weight
(
Kannan
et
al.,
1994).
Levengood
et
al.
(
1999)
analyzed
44
samples
of
Canadian
goose
meat
collected
in
northeastern
Illinois
during
1994
for
pesticide
residues.
Dieldrin
was
detected
in
16%
of
the
baked
skinless
samples
at
concentrations
ranging
from
0.004
to
0.011
mg/
kg,
and
in
7%
of
the
samples
baked
with
the
skin
and
overlying
adipose
tissue
at
concentrations
of
0.005
to
0.010
mg/
kg.
Dieldrin
residue
levels
reported
in
this
study
were
below
FDA
residue
limits
of
0.30
mg/
kg
(
Dey
and
Manzoor,
1997).

Milk
and
milk
products
are
additional
sources
of
dieldrin
in
the
diet.
During
1990
and
1991,
milk
samples
were
collected
from
63
metropolitan
locations
throughout
the
United
States,
as
part
of
the
EPA's
Pasteurized
Milk
Program.
Dieldrin
was
detected
in
21.1%
of
806
composited
milk
samples
at
concentrations
ranging
from
0.0005
mg/
kg
(
detection
limit)
to
0.002
mg/
kg
(
Trotter
and
Dickerson,
1993).
FDA
Total
Diet
Study
results
from
1985
through
1991
reported
mean
dieldrin
concentrations
in
whole
milk,
2%
milk,
evaporated
canned
milk,
and
chocolate
milk
samples
of
0.0003
mg/
kg,
0.0003
mg/
kg,
0.0008
mg/
kg,
and
0.0014
mg/
kg,
respectively
(
KAN­
DO
Office
and
Pesticides
Team,
1995).
Maximum
dieldrin
concentrations
detected
in
vitamin
D
milk
and
plain
milk
samples
as
part
of
the
FDA
Regulatory
Monitoring
were
0.03
mg/
kg
and
1
mg/
kg,
respectively.
The
maximum
residue
found
in
whole
milk
(
1
mg/
kg)
was
above
the
EPA
milk
tolerance
of
0.30
ppm
(
0.30
mg/
kg)
(
Yess
et
al.,
1993).

Dingle
et
al.
(
1989)
found
dieldrin
to
persist
in
milk
butterfat,
with
a
half­
life
in
butter
of
approximately
9
weeks.
Ultra­
pasteurized
heavy
cream
and
cow
milk
samples
purchased
in
Binghamton,
New
York,
in
1986
had
dieldrin
levels
of
0.006
mg/
kg
and
0.003
mg/
kg,
respectively
(
Schecter
et
al.,
1989).
5­
4
Aldrin/
Dieldrin
 
February
2003
Infant
foods
analyzed
during
the
FDA's
Market
Basket
Survey
from
1981
through
1992
contained
mean
dieldrin
residues
ranging
from
0.003
to
0.0051
mg/
kg
(
KAN­
DO
Office
and
Pesticides
Team,
1995).
Maximum
dieldrin
concentrations
detected
in
infant
foods
sampled
during
the
1985
to
1991
sampling
period
as
part
of
the
FDA's
Total
Diet
Study
were
0.002
mg/
kg.
Adult
foods
eaten
by
infants
and
children
also
analyzed
as
part
of
the
FDA
Total
Diet
Study
and
Regulatory
Monitoring
programs
(
from
1985
through
1991)
detected
dieldrin
in
creamy
peanut
butter,
pears,
and
one
imported
orange
at
maximum
concentrations
of
0.003,
0.0005,
and
0.01
mg/
kg,
respectively
(
Yess
et
al.,
1993).

Because
many
infants
receive
human
breast
milk,
their
dieldrin
intakes
may
be
closely
related
to
its
concentration
in
human
breast
milk.
Current
data
regarding
the
levels
of
dieldrin
in
human
breast
milk
in
the
United
States
were
not
located.
However,
data
from
several
older
studies
are
available.
Dieldrin
was
found
in
the
breast
milk
of
80.8%
of
1,436
nursing
women
sampled
in
1980,
with
a
mean
fat­
adjusted
residue
level
of
0.164
mg/
kg
(
Savage
et
al.,
1981).
Additional
studies
of
nursing
mothers
in
Hawaii
(
Takei
et
al.,
1983)
and
in
Mississippi
and
Arkansas
(
Strassman
and
Kutz,
1977)
found
dieldrin
residues
in
breast
milk
at
mean
concentrations
of
1.3
ppb
(
0.0013
mg/
kg)
and
4
ppb
(
0.004
mg/
kg),
respectively.
Breast
milk
collected
from
Canadian
provinces
during
1986
contained
an
average
dieldrin
concentration
of
4.6
×
10­
5
ppm
(
4.6
×
10­
5
mg/
kg)
(
Mes
et
al.,
1993).

Intake
from
Non­
Fish
Food
Items
Aldrin
The
mean
aldrin
concentration
detected
in
domestic
food
items
during
1981
to
1992
was
0.0016
mg/
kg
(
KAN­
DO
Office
and
Pesticides
Team,
1995).
Based
on
this
concentration,
a
70
kg
adult
with
a
food
intake
rate
of
1.305
kg/
day
(
USEPA,
1988)
would
have
an
average
daily
aldrin
intake
of
3.0
×
10­
5
mg/
kg­
day.
At
the
same
concentration,
the
average
daily
aldrin
intake
for
a
10
kg
child
would
be
1.3
×
10­
4
mg/
kg­
day,
assuming
a
food
intake
rate
of
0.84
kg/
day
(
USEPA,
1988).
These
intakes
are
based
on
the
mean
aldrin
concentrations
of
positive
samples.
Food
samples
where
aldrin
was
not
detected
are
not
included
in
the
average.
Thus
these
estimated
daily
intakes
of
aldrin
from
food
overestimate
the
true
mean
for
the
general
population.
ATSDR
(
2000)
reports
average
aldrin
intakes
to
be
approximately
<
0.001
:
g/
kg/
day
(<
1.0
×
10­
6
mg/
kg­
day).

Dieldrin
Dieldrin
was
detected
more
frequently
in
food
items
than
aldrin.
The
mean
dieldrin
concentration
in
food
items
analyzed
during
FDA
Market
Basket
Study
1981­
1992
(
KAN­
DO
Office
and
Pesticides
Team,
1995)
was
0.0015
:
g/
g.
Based
on
this
average
concentration,
a
70
kg
adult
with
a
food
intake
rate
of
1.305
kg/
day
(
USEPA,
1988)
would
have
an
average
daily
dieldrin
intake
of
2.8
×
10­
5
mg/
kg­
day.
A
10
kg
child,
with
a
food
intake
rate
of
0.84
kg/
day
(
USEPA,
1988)
would
have
a
daily
dieldrin
intake
rate
of
1.3
×
10­
4
mg/
kg­
day.
These
estimates
are
based
on
the
mean
of
dieldrin
concentrations
in
positive
samples
and
does
not
incorporate
food
samples
without
detectable
levels
of
dieldrin
into
the
average.
Thus,
these
estimates
will
overestimate
the
typical
dieldrin
intakes
experienced
by
the
general
population.
Additional
5­
5
Aldrin/
Dieldrin
 
February
2003
studies
have
estimated
dietary
intakes
of
dieldrin.
MacIntosh
et
al.
(
1996)
estimated
daily
dieldrin
dietary
intakes
for
adults
to
range
from
2
×
10­
5
to
4
×
10­
3
mg/
day,
with
a
mean
of
approximately
5
×
10­
4
mg/
day.
These
estimates
are
based
on
mean
dieldrin
concentrations
reported
for
234
ready­
to­
eat
food
items
from
the
FDA's
Total
Diet
Study
during
1986
through
1991
and
approximately
117,000
food
consumption
surveys
from
the
Nurses'
Health
Study
and
the
Health
Professionals/
Follow­
up
Study.
Gunderson
(
1988)
estimated
daily
dieldrin
intakes
for
adults
to
be
7
×
10­
6
to
8
×
10­
6
mg/
kg­
day
during
1982
to
1984.

Rogan
and
Ragan
(
1994)
estimated
a
high­
end
average
daily
intake
(
90th
percentile)
of
dieldrin
for
infants
through
breast
milk
in
the
United
States
to
be
3.6
×
10­
6
mg/
kg­
day.
This
estimate
is
based
on
dieldrin
concentrations
in
breast
milk
of
0.10
ppm
fat
(
Savage
et
al.,
1984),
and
daily
intakes
of
700
g
of
breast
milk
(
2.5%
fat)
per
day
for
9
months.

Concentrations
in
Fish
and
Shellfish
Aldrin
Two
studies
were
located
that
reported
aldrin
concentrations
in
fish
and
shellfish.
Murray
and
Beck
(
1990)
analyzed
shrimp
(
Penaeus
setiferus
and
Penaeus
aztecus)
collected
from
30
stations
along
the
Calcasieu
River
Basin
in
an
industrial
area
of
Louisiana
during
1985
to
1986.
Aldrin
was
detected
in
shrimp
samples
from
7
of
the
30
stations,
at
concentrations
ranging
from
0.01
to
0.12
:
g/
g
(
0.01
to
0.12
mg/
kg).

In
another
study,
Kannan
et
al.
(
1994)
reported
aldrin
concentrations
for
fish
and
shellfish
samples
collected
from
various
metropolitan
locations
in
Australia,
Papua
New
Guinea,
and
the
Solomon
Islands
during
1990.
Mean
aldrin
concentrations
were
2.1
×
10­
3,
4.5
×
10­
4,
and
7.7
×
10­
4
mg/
kg
(
wet
weight)
for
oyster,
mudcrab,
and
fish
samples,
respectively.

Dieldrin
Several
studies
have
reported
dieldrin
residues
in
fish
and
shellfish.
Bottom
feeding
and
game
fish
sampled
from
400
sites
throughout
the
United
States
between
1986
and
1989
as
part
of
the
National
Study
of
Chemical
Residues
in
Fish
Survey
contained
mean
dieldrin
concentrations
of
28.1
ng/
g
(
0.0281
mg/
kg).
Of
the
119
total
fish
species
sampled,
the
5
most
frequently
sampled
fish
species
and
their
respective
dieldrin
concentrations
were
as
follows:
Carp
(
0.0448
mg/
kg),
White
Sucker
(
0.0228
mg/
kg)
and
Channel
Catfish
(
0.0154
mg/
kg),
Largemouth
Bass
(
0.005
mg/
kg),
Smallmouth
Bass
(
0.00234
mg/
kg),
and
Walleye
(
0.00373
mg/
kg)
(
Kuehl
et
al.,
1994).

Dieldrin
concentrations
analyzed
in
11
species
of
fish
in
the
Great
Lakes
ranged
from
0.24
to
41.2
ng/
g
wet
weight
(
0.00024
to
0.041
mg/
kg
wet
weight).
The
highest
dieldrin
concentrations
were
detected
in
carp
(
0.040
mg/
kg),
trout
(
0.041
mg/
kg),
and
eel
(
0.031
mg/
kg).
Bullhead
(
0.00024
mg/
kg)
and
perch
(
0.00098
mg/
kg)
contained
the
lowest
dieldrin
concentrations
(
Newsome
and
Andrews,
1993).
Walleye
and
white
bass
samples
(
skin
on)
contained
mean
dieldrin
concentrations
ranging
from
0.006
to
0.009
mg/
kg
wet
weight
and
0.011
mg/
kg
wet
weight,
respectively,
in
raw
samples
collected
from
the
Great
Lakes
during
April
and
5­
6
Aldrin/
Dieldrin
 
February
2003
July
1991.
Pan
frying
white
bass
samples
(
skin
removed)
reduced
dieldrin
concentrations
on
average
by
34.8%.
Dieldrin
loss
from
deep
fat
frying
(
skin
and
muscle)
walleye
samples
was
26.4%
(
Zabik
et
al.,
1995).

Fairey
et
al.
(
1997)
measured
pesticide
concentrations
in
fish
species
commonly
caught
by
anglers
from
16
areas
throughout
the
San
Francisco
Bay
during
1994.
Dieldrin
was
detected
in
six
of
the
seven
species
of
fish
analyzed.
As
listed
in
Table
5­
2,
dieldrin
concentrations
in
the
seven
fish
species
ranged
from
non­
detectable
to
4.2
ng/
g
(
0.0042
mg/
kg)
wet
weight.
Concentrations
were
proportional
to
fish
lipid
content.
White
croaker
fish
samples
had
the
highest
dieldrin
levels,
and
also
the
highest
lipid
content.
Fish
species
with
lower
lipid
contents
(
sharks
and
halibut)
had
the
lowest
dieldrin
concentrations.

Blynn
et
al.
(
1994)
analyzed
two
composited
filet
samples
from
three
stations
in
Pennekamp
Coral
Reef
State
Park
and
Key
Largo
National
Marine
Sanctuary
for
pesticide
residues
during
September
1992.
None
of
the
filet
samples
contained
dieldrin
concentrations
above
the
detection
limit
of
0.001
mg/
kg.

Shrimp
(
Penaeus
setiferus
and
Penaeus
aztecus)
samples
collected
from
21
of
30
stations
along
the
Calcasieu
River
Basin
in
an
industrial
area
of
Louisiana
during
1985
to
1986
contained
mean
dieldrin
concentrations
of
1.57
:
g/
g
(
1.57
mg/
kg).
Dieldrin
concentrations
ranged
from
0.05
to
9.47
:
g/
g
(
0.05
to
9.47
mg/
kg)
(
Murray
and
Beck,
1990).

Kannan
et
al.
(
1994)
reported
dieldrin
levels
in
fish
and
shellfish
samples
collected
from
various
metropolitan
locations
in
Australia,
Papua
New
Guinea,
and
the
Solomon
Islands
during
1990.
Mean
dieldrin
concentrations
were
7.3
×
10­
4,
3.2
×
10­
4,
and
9.5
×
10­
3
mg/
kg
(
wet
weight)
for
oyster,
mudcrab,
and
fish
samples,
respectively.

Table
5­
2.
Aldrin
Concentrations
in
San
Francisco
Bay
Area
Fish
in
19941
Fish
Species
Dieldrin
Concentration
Number
of
Fish
Sampled
White
Croaker
1.1
x
10­
3
to
4.2
x
10­
3
125
Striped
Bass
1.1
x
10­
3
to
3.0
x
10­
3
16
Shiner
Surf
Perch
ND2
to
2.59
x
10­
3
160
Leopard
Shark
ND
to
6.1
x
10­
4
14
Brown
Smoothhound
Shark
ND­
0.000341
21
Sturgeon
3.1
x
10­
3
3
Halibut
ND
3
1
Source:
Fairley
et
al.
(
1997).

2
ND:
Not
Detected
(
detection
limit
not
reported).
5­
7
Aldrin/
Dieldrin
 
February
2003
Intake
from
Fish
and
Shellfish
Aldrin
Only
one
study
was
located
that
reported
aldrin
concentrations
in
fish
and
shellfish
(
Murray
and
Beck,
1990).
Shrimp
samples
collected
from
an
industrial
area
of
Louisiana
contained
aldrin
concentrations
ranging
from
0.01
to
0.12
mg/
kg.
Based
on
these
concentrations,
and
an
average
daily
intake
of
20.1
g/
day
(
USEPA,
1997),
a
70
kg
adult
would
have
an
average
daily
intake
of
2.9
×
10­
6
to
3.5
×
10­
5
mg/
kg­
day.
A
10
kg
child
with
a
daily
intake
rate
of
4.0
g/
day
(
USEPA,
1997)
would
have
a
daily
aldrin
intake
of
4.0
×
10­
6
to
4.8
×
10­
5
mg/
kg­
day.
These
intakes
are
based
on
aldrin
concentrations
in
fish
from
an
industrial
area,
which
may
be
higher
than
typical
aldrin
levels
in
fish.
Thus,
these
estimated
aldrin
intakes
may
not
be
representative
of
general
population
exposures
to
aldrin
in
fish.

Dieldrin
Assuming
an
average
concentration
of
dieldrin
in
fish
of
0.0281
mg/
kg
(
Kuehl
et
al.,
1994),
and
a
daily
in
take
of
20.1
g/
day
(
USEPA,
1997),
a
70
kg
adult
would
have
an
average
dieldrin
daily
intake
of
8.0
×
10­
6
mg/
kg­
day.
A
10
kg
child
exposed
to
the
same
concentrations
would
have
a
daily
dieldrin
intake
of
1.1
×
10­
5,
based
on
a
daily
intakes
of
4.0
g/
day
(
USEPA,
1997).
Ahmed
et
al.
(
1993)
estimated
dietary
exposures
to
dieldrin
from
American
finfish
to
be
4.9
×
10­
7
mg/
kg­
day,
based
on
FDA
surveillance
data
collected
from
1984
to
1988.

5.1.2
Exposures
of
Subpopulations
Persons
working
with
or
living
in
areas
utilizing
aldrin
and
dieldrin
may
potentially
have
higher
concentrations
of
these
pesticides
in
their
diets
(
Melnyk
et
al.,
1997).

Concentrations
in
Food
Items
Aldrin
Additional
information
on
concentrations
of
aldrin
in
non­
fish
food
items
and
fish/
shellfish
or
on
intakes
of
aldrin
by
subpopulations
were
not
obtained
in
the
available
literature.

Dieldrin
One
study
was
located
that
analyzed
dieldrin
concentrations
in
the
diets
of
farmers
(
Melnyk
et
al.,
1997).
Food
samples
from
six
farms
in
Iowa
and
North
Carolina
were
analyzed
during
both
a
pesticide
application
and
non­
application
period
as
part
of
a
pilot
study
to
evaluate
pesticide
exposures
of
farmers
and
their
families.
Food
and
beverage
samples
at
one
of
the
six
farms
had
dieldrin
concentrations
ranging
from
11
to
28
ppb
(
0.011
to
0.028
mg/
kg).
Food
samples
collected
during
the
non­
application
period
had
higher
dieldrin
concentrations
than
those
collected
during
the
application
period
of
28
ppb
(
0.028
mg/
kg)
and
15
ppb
(
0.015
mg/
kg),
respectively.
Dieldrin
was
not
detected
in
beverages
collected
during
application
periods,
5­
8
Aldrin/
Dieldrin
 
February
2003
whereas
beverages
sampled
during
the
non­
application
period
contained
11
ppb
(
0.011
mg/
kg)
dieldrin.
Previous
aldrin
use
at
the
farm,
the
presence
of
dieldrin
in
milk
(
0.008
to
0.015
mg/
kg)
from
area
dairy
farms
(
Bond
et
al.,
1993),
and
the
general
persistence
of
dieldrin
in
the
Midwest
(
MacMonegle
et
al.,
1984)
may
all
contribute
to
the
high
dieldrin
concentrations
detected
in
food
items
at
this
farm.
Dieldrin
was
not
detected
in
food
and
beverage
samples
from
the
other
five
farms
in
the
pilot
study.
Details
on
the
types
of
foods
(
e.
g.,
fish
and
non­
fish
food
items)
analyzed
in
the
pilot
study
were
not
provided.

Intake
from
Food
Items
Aldrin
Additional
information
on
concentrations
of
aldrin
in
non­
fish
food
items
and
fish/
shell
fish
or
on
intakes
of
aldrin
by
subpopulations
were
not
obtained
in
the
available
literature.
Thus,
intakes
of
aldrin
by
subpopulations
were
not
calculated.

Dieldrin
Melnyk
et
al.
(
1997)
detected
dieldrin
in
food
and
beverages
in
the
diets
of
farmers
during
a
pilot
study
of
farms
in
Iowa
and
North
Carolina.
Dieldrin
was
detected
in
food
items
at
one
of
the
six
farms
with
a
history
of
aldrin
usage
analyzed
in
the
study.
Mean
dieldrin
concentrations
in
food
items
were
28
ppb
(
0.028
mg/
kg)
and
15
ppb
(
0.015
mg/
kg)
for
nonapplication
and
application
periods,
respectively.
Based
on
these
concentrations
(
0.015
to
0.028
mg/
kg)
and
an
intake
of
1.305
kg/
day
(
USEPA,
1988),
a
70
kg
adult
worker
would
have
an
average
daily
dieldrin
intake
ranging
from
2.8
×
10­
4
to
5.2
×
10­
4
mg/
kg­
day.

5.2
Exposure
from
Air
Aldrin
and
dieldrin
have
both
been
used
for
pest
control
in
agriculture
and
as
termiticides.
Agricultural
uses
of
aldrin
and
dieldrin
were
cancelled
in
1974
and
their
use
as
a
termiticide
cancelled
in
1987.
Aldrin
and
dieldrin
may
enter
the
atmosphere
through
mechanisms
such
as
spray
drift
during
application,
water
evaporation,
and
dispersion
and
suspension
of
particulates
or
soils
to
which
the
compounds
are
absorbed
(
ATSDR,
2000).

5.2.1
Exposures
of
the
General
Population
Concentrations
in
Air
Aldrin
Current
data
on
ambient
concentrations
of
aldrin
in
air
were
not
located
in
the
available
literature.
However,
from
1970
to
1972
Kutz
et
al.
(
1976)
analyzed
2,479
air
samples
from
16
states.
Aldrin
was
detected
in
13.5%
of
the
samples
with
a
mean
of
3
×
10­
5
ppb
(
4
×
10­
7
mg/
m3).
Ambient
concentrations
reported
by
this
study
are
likely
higher
than
current
ambient
aldrin
levels,
as
it
was
conducted
prior
to
the
cancellation
of
all
uses
of
adrin
and
dieldrin.
5­
9
Aldrin/
Dieldrin
 
February
2003
Several
studies
have
measured
indoor
air
concentrations
of
aldrin,
as
the
potential
for
higher
exposure
rates
may
occur
for
segments
of
the
population
residing
in
homes
using
this
chemical
for
termite
control
(
Dobbs
and
Williams,
1983).

A
pilot
study
of
non­
occupational
exposures
to
pesticides
for
the
general
population
from
ambient
air
inside
and
outside
the
home
was
conducted
in
nine
homes
during
1985.
Indoor
and
outdoor
air,
as
well
as
personal
air
monitors,
were
sampled
over
24­
hour
periods.
Aldrin
was
detected
in
indoor
air
at
six
of
the
nine
households;
outdoors
at
four
of
the
nine
households;
and
in
three
of
the
nine
personal
monitors.
In
one
designated
high­
pesticide­
use
household,
aldrin
was
detected
in
the
indoor
air
at
average
concentrations
of
0.004
ppb
(
5.8
×
10­
5
mg/
m3).
Neither
compound
was
detected
in
the
outdoor
air
immediately
adjacent
to
the
home
and
concentrations
detected
with
personal
air
monitors
were
half
of
the
concentrations
reported
for
indoor
air
samples
(
Lewis
et
al.,
1988).

Indoor
air
concentrations
of
aldrin
were
monitored
on
each
level
of
a
two­
story
home
in
Bloomington,
Indiana,
(
Wallace
et
al.,
1996)
identified
in
a
previous
study
(
Anderson
and
Hites,
1988)
as
having
elevated
concentrations
of
these
chemicals.
Aldrin
had
been
poured
into
the
void
spaces
of
the
foundation
blocks
during
its
construction
in
1985
for
termite
control.
Between
September
1987
and
April
1995,
aldrin
concentrations
had
decreased
from
5,000
ng/
m3
to
12
ng/
m3
(
5
×
10­
3
to
1.2
×
10­
5
mg/
m3)
in
the
basement,
and
from
300
ng/
m3
to
2
ng/
m3
(
3
×
10­
4
to
2
×
10­
6
mg/
m3)
in
the
living
area.

Dieldrin
Several
studies
have
measured
dieldrin
in
ambient
air.
Kutz
et
al.
(
1976)
analyzed
2,479
air
samples
from
16
states
from
1970
to
1972.
Dieldrin
was
detected
in
94%
of
samples
with
a
mean
of
1
x
10­
4
ppb
(
1.6
×
10­
6
mg/
m3).

In
another
study,
dieldrin
was
detected
at
an
average
concentration
of
5.1
×
10­
6
ppb
(
8.0
×
10­
8
mg/
m3)
in
ambient
air
over
College
Station,
Texas,
during
1979
through
1980
(
Atlas
and
Giam,
1988).

Several
studies
have
measured
indoor
air
concentrations
of
dieldrin.
One
study
reported
dieldrin
concentrations
in
indoor
air
for
homes
1
to
10
years
after
the
termiticide
treatment
ranging
from
0.002
to
0.17
ppb
(
3.16
×
10­
5
to
1.98
×
10­
3
mg/
m3)
in
roof
voids,
and
from
0.0006
to
0.03
ppb
(
9.49
×
10­
6
to
4.75
×
10­
4
mg/
m3)
in
living
rooms,
bedrooms,
and
all
interior
areas
(
Dobbs
and
Williams,
1983).

A
pilot
study
of
non­
occupational
exposures
to
pesticides
for
the
general
population
from
ambient
air
inside
and
outside
the
home
was
conducted
in
nine
homes
during
1985.
Indoor
and
outdoor
air,
as
well
as
personal
air
monitors,
were
sampled
over
24­
hour
periods.
Dieldrin
was
detected
in
indoor
air
at
five
of
the
nine
households;
outdoors
at
four
of
the
nine
households;
and
by
personal
monitors
for
five
out
of
nine
individuals.
In
one
designated
high­
pesticide­
use
household,
dieldrin
was
detected
in
the
indoor
air
at
average
concentrations
of
0.002
ppb
(
3.8
×
10­
5
mg/
m3).
Neither
compound
was
detected
in
the
outdoor
air
immediately
adjacent
to
5­
10
Aldrin/
Dieldrin
 
February
2003
the
home
and
concentrations
detected
with
personal
air
monitors
were
one­
third
the
concentrations
for
ambient
indoor
air
(
Lewis
et
al.,
1988).

Indoor
air
concentrations
of
dieldrin
were
monitored
on
each
level
of
a
two­
story
home
in
Bloomington,
Indiana,
identified
in
a
previous
study
(
Anderson
and
Hites,
1988)
as
having
elevated
concentrations
of
these
chemicals.
Aldrin
had
been
poured
into
the
void
spaces
of
the
foundation
blocks
during
its
construction
in
1985
for
termite
control.
Between
September
1987
and
April
1995,
dieldrin
concentrations
fell
from
28
ng/
m3
to
20
ng/
m3
(
2.8
×
10­
5
to
2.0
×
10­
5
mg/
m3)
in
the
basement,
and
from
7
ng/
m3
to
3
ng/
m3
(
7
×
10­
6
to
3
×
10­
6
mg/
m3)
in
the
living
area
(
Wallace
et
al.,
1996).

Indoor
air
samples
were
collected
as
part
of
a
pilot
study
in
Raleigh,
North
Carolina,
to
characterize
pesticide
exposures
of
children.
Samples
were
collected
at
2
different
heights
(
12.5
cm
and
75
cm)
from
the
living
rooms
of
8
homes,
over
a
24­
hour
period.
Dieldrin
was
detected
in
indoor
air
samples
at
four
of
the
eight
homes,
at
a
mean
concentration
of
0.01
:
g/
m3
(
1
×
10­
5
mg/
m3),
and
at
a
maximum
concentration
of
0.02
:
g/
m3
(
2
×
10­
5
mg/
m3)
(
Lewis
et
al.,
1994).

Intake
from
Air
Aldrin
Intake
of
aldrin
from
air
was
estimated
based
on
the
mean
ambient
air
concentration
reported
by
Kutz
et
al.
(
1976)
from
1970
to
1972
of
4
×
10­
7
mg/
m3.
Assuming
an
inhalation
rate
of
20
m3/
day
(
USEPA,
1988),
the
average
estimated
daily
intake
of
aldrin
for
a
70
kg
adult
would
be
1.1
×
10­
7
mg/
kg­
day.
The
estimated
average
daily
intake
of
aldrin
for
a
10
kg
child
is
6.0
×
10­
7
mg/
kg­
day,
based
on
an
inhalation
rate
of
15
m3/
day
(
USEPA,
1988).
This
ambient
concentration
of
aldrin
was
measured
prior
to
the
cancellation
of
all
uses
of
aldrin
and
dieldrin.
Thus,
these
estimated
daily
intakes
of
aldrin
from
air
will
overestimate
general
population
exposures
from
air.

Dieldrin
The
mean
dieldrin
concentration
reported
for
ambient
air
from
1970
to
1972
is
1.6
×
10­
6
mg/
m3
(
Kutz
et
al.,
1976).
Assuming
an
inhalation
rate
of
20
m3/
day
(
USEPA,
1988),
the
average
estimated
daily
intake
dieldrin
for
a
70
kg
adult
would
be
4.6
×
10­
7
mg/
kg­
day.
The
estimated
average
daily
intake
of
dieldrin
in
air
for
a
10
kg
child
is
2.4
×
10­
6
mg/
kg­
day,
based
on
an
inhalation
rate
of
15
m3/
day
(
USEPA,
1988).
These
estimated
daily
intakes
will
overestimate
general
population
exposures
to
dieldrin
from
air,
as
they
are
based
on
ambient
air
concentrations
reported
prior
to
the
cancellation
of
all
uses
of
aldrin
and
dieldrin.

Higher
intakes
of
aldrin
and
dieldrin
may
be
expected
for
populations
living
in
homes
using
these
chemicals
for
termite
control
(
ATSDR,
2000).
5­
11
Aldrin/
Dieldrin
 
February
2003
5.2.2
Exposures
of
Subpopulations
Persons
involved
in
the
manufacturing
or
application
of
aldrin
or
dieldrin
may
potentially
be
exposed
to
these
chemicals
in
air.
However,
data
on
workplace
or
post­
application
concentrations
of
aldrin
or
dieldrin
in
air,
or
intakes
of
these
chemicals
by
workers
were
not
available
from
the
retrieved
literature.

5.3
Exposure
from
Soil
Aldrin
and
dieldrin
were
used
as
pesticides,
until
their
registrations
were
cancelled
in
1974.
Although
aldrin
was
applied
more
frequently
to
soils,
dieldrin
is
found
more
often
and
in
higher
concentrations
than
aldrin
residues
(
ATSDR,
2000).

5.3.1
Exposures
of
the
General
Population
Concentrations
in
Soil
Aldrin
Data
on
aldrin
in
residential
soils
were
not
located
in
the
available
literature.
Based
on
the
rapid
conversion
of
aldrin
to
dieldrin
in
soils
(
ATSDR,
2000),
the
general
population
is
more
likely
to
be
exposed
to
dieldrin
than
aldrin
from
soil.

Dieldrin
The
National
Soils
Monitoring
Program
(
Kutz
et
al.,
1976)
detected
dieldrin
in
soils
throughout
24
states
at
mean
concentrations
ranging
from
1
to
49
ppb
(
0.001
to
0.049
mg/
kg).

Pesticides
may
accumulate
in
carpets
from
indoor
treatment
and
the
tracking
in
of
outdoor
soils,
thus
contributing
to
residential
exposures
(
Lewis
et
al.,
1994).
A
composite
sample
of
the
dust
from
four
Seattle
homes
collected
during
1988
to
1989
contained
1.1
mg/
kg
dieldrin,
although
none
of
the
homeowners
could
remember
using
the
pesticide
(
Roberts
and
Camann,
1989).

Lewis
et
al.
(
1994)
analyzed
house
dust
and
soil
samples
from
nine
homes
in
North
Carolina,
varying
in
pesticide
use,
as
part
of
a
pilot
study
to
evaluate
monitoring
methods
used
to
assess
exposures
to
children.
House
dust
samples
were
collected
by
taking
40
passes
over
a
3800
cm2
carpet
areas
of
the
homes
with
a
HVS3
vacuum
system.
The
mean
dieldrin
concentration
was
0.29
mg/
kg
(
0.12
:
g/
m2),
with
a
maximum
concentration
of
1.0
mg/
kg
(
0.38
:
g/
m2).
Entryway
soil
samples,
collected
from
outside
the
doorway
most
frequently
used,
had
mean
dieldrin
concentrations
of
0.07
mg/
kg,
and
a
maximum
of
0.19
mg/
kg
at
four
of
the
nine
homes
sampled.
Soils
(
up
to
0.5
mm
in
depth)
collected
from
childrens'
play
areas
contained
mean
dieldrin
concentrations
of
0.03
mg/
kg
and
a
maximum
concentration
of
0.09
mg/
kg.
Higher
dieldrin
levels
were
found
in
soils
from
primary
walkways
of
0.26
mg/
kg
(
mean)
and
a
maximum
concentration
of
0.54
mg/
kg
(
Lewis
et
al.,
1994).
5­
12
Aldrin/
Dieldrin
 
February
2003
Intake
from
Soil
Aldrin
Data
on
aldrin
levels
in
residential
soils
were
not
located
in
the
available
literature.
Thus,
average
daily
intakes
of
aldrin
by
the
general
population
from
soil
could
not
be
estimated.
The
use
of
aldrin
as
a
pesticide
was
cancelled
in
1974.
Based
on
its
cancellation
and
its
rapid
conversion
to
dieldrin
in
the
environment
(
ATSDR,
2000),
it
is
assumed
that
the
general
population
is
more
likely
to
be
exposed
to
dieldrin
than
aldrin
in
soils.

Dieldrin
Dieldrin
has
been
detected
in
both
residential
soils
(
Lewis
et
al.,
1994)
and
house
dust
(
Roberts
and
Camann,
1989)
samples.
Mean
dieldrin
concentrations
ranged
from
0.03
to
1.1
mg/
kg.
Based
on
this
range
of
concentrations,
and
a
daily
intake
of
50
mg/
day
(
USEPA,
1997)
for
a
70
kg
adult,
the
total
daily
intake
of
dieldrin
through
soil
ranges
from
2.1
×
10­
8
mg/
kg­
day
to
7.9
×
10­
7
mg/
kg­
day.
For
a
10
kg
child
exposed
to
the
same
soil
concentrations,
at
an
intake
rate
of
100
mg/
day
(
USEPA,
1997),
the
total
daily
dieldrin
intake
would
be
3.0
×
10­
7
mg/
kg­
day
to
1.1
×
10­
5
mg/
kg­
day.

5.3.2
Exposures
of
Subpopulations
Persons
involved
in
the
manufacture,
handling,
or
application
of
aldrin
and
dieldrin
may
potentially
have
higher
exposures
to
these
chemicals
from
soil
through
incidental
ingestion.

Concentrations
in
Soil
Aldrin
Data
on
aldrin
concentrations
in
agricultural
soils
in
the
Unites
States
were
not
located
in
the
available
literature.
However,
one
study
reported
aldrin
in
soil
samples
collected
from
agricultural
fields
in
Farrukhabad,
India
from
1991
to
1992.
Surface
soil
samples
(
0
to
15
cm)
contained
aldrin
concentrations
ranging
from
0.001
to
0.010
mg/
kg,
with
means
ranging
from
0.001
to
0.004
mg/
kg.
Subsurface
(
15
to
30
cm)
concentrations
ranged
from
0.001
to
0.014
mg/
kg,
with
means
of
0.001
to
0.006
mg/
kg
(
Agnihotri
et
al.,
1996).

Dieldrin
Several
studies
have
evaluated
dieldrin
residues
in
agricultural
soils.
Aigner
et
al.
(
1998)
sampled
38
agricultural
soils
from
Ohio,
Pennsylvania,
Indiana,
and
Illinois
during
1995
and
1996
for
pesticide
residues.
Dieldrin
was
detected
in
21
of
38
soils
at
concentrations
ranging
from
0.12
to
71
ng/
g
(
0.00012
to
0.071
mg/
kg).
One
soil
sample
from
Ohio
had
considerably
higher
dieldrin
concentrations
than
the
other
soil
samples
with
residues
of
4.25
mg/
kg.
This
soil
sample
contained
the
highest
concentrations
of
all
individual
pesticides
analyzed.
Samples
from
two
garden
soils
contained
4.39
ng/
g
(
0.0044
mg/
kg)
and
3.47
ng/
g
(
0.0035
mg/
kg)
of
dieldrin.
5­
13
Aldrin/
Dieldrin
 
February
2003
Harner
et
al.
(
1999)
reported
dieldrin
concentrations
ranging
from
<
0.02
to
23.9
ng/
g
dry
weight
(<
0.00002
to
0.024
mg/
kg),
and
a
mean
of
0.0049
mg/
kg
for
36
agricultural
soils
surveyed
throughout
Alabama.

The
persistence
of
dieldrin
in
agricultural
fields
is
demonstrated
by
a
monitoring
survey
conducted
in
and
around
cotton
fields
in
four
counties
in
Alabama
between
1972
and
1974.
Although
aldrin
or
dieldrin
had
not
been
reportedly
used
by
cotton
farmers
"
for
several
years,"
dieldrin
was
found
to
be
present
in
50%
of
the
soil
samples,
at
concentrations
ranging
from
0.007
to
0.040
mg/
kg
(
Elliott,
1975).

Intakes
from
Soil
Aldrin
Data
on
aldrin
concentrations
in
agricultural
fields
in
the
United
States
were
not
located
in
the
available
literature.
Although
one
study
(
Agnihotri
et
al.,
1996)
was
located
that
detected
aldrin
levels
in
agricultural
soils,
it
reported
residues
for
soils
in
India.
This
study
is
not
representative
of
exposures
to
aldrin
from
agricultural
soils
that
may
occur
in
the
United
States.
The
uses
of
aldrin
as
a
pesticide
and
termiticide
have
been
cancelled
since
1974
and
1987,
respectively.
Based
on
the
cancellations
of
its
uses
in
the
United
States
and
the
rapid
conversion
of
aldrin
to
dieldrin
in
the
environment
(
ATSDR,
2000),
subpopulations
are
more
likely
to
be
exposed
to
dieldrin
than
aldrin
in
soils.

Dieldrin
Several
studies
have
reported
dieldrin
concentrations
in
agricultural
soils
of
the
United
States.
These
concentrations
range
from
<
0.00002
to
0.071
mg/
kg
(
Harner
et
al.,
1999
and
Aigner
et
al.,
1998).
Based
on
these
concentrations
and
an
intake
rate
of
480
mg/
day
(
USEPA,
1997),
for
a
contact
intensive
worker,
the
average
daily
intake
of
dieldrin
from
soil
for
a
70
kg
adult
worker
would
range
from
1.4
×
10­
10
to
2.9
×
10­
5
mg/
kg­
day.
A
high­
end
estimate
of
potential
subpopulation
exposures
to
dieldrin
in
agricultural
soils
can
be
determined
based
on
the
highest
concentration
reported
by
Aigner
et
al.
(
1998)
of
4.25
mg/
kg.
At
this
concentration
and
an
intake
rate
of
480
mg/
day
(
USEPA,
1997),
a
70
kg
adult,
contact
intensive
worker
would
have
an
average
daily
dieldrin
intake
of
2.9
×
10­
5
mg/
kg­
day.

5.4
Other
Residential
Exposures
(
Not
Drinking
Water
Related)

Aldrin
and
dieldrin
residues
have
been
reported
in
rainfall
and
carpet.
Dieldrin
has
additionally
been
detected
in
sediments.

Aldrin
Aldrin
was
detected
in
rainfall
collected
from
the
Great
Lakes
Basin
during
1986,
approximately
10
years
after
aldrin
and
dieldrin
use
was
restricted.
Aldrin
was
present
in
wet
precipitation
at
three
of
four
sampling
sites
located
around
the
basin,
in
6.7%
of
the
samples
collected
at
a
mean
concentrations
ranging
from
0.01
ng/
L
(
1
×
10­
5
ppb)
to
0.24
ng/
L
(
2.4
×
10­
4
5­
14
Aldrin/
Dieldrin
 
February
2003
ppb).
The
highest
aldrin
concentrations
were
found
in
samples
collected
at
Pelee
Island
at
the
western
end
of
Lake
Erie
at
a
maximum
concentration
of
3.4
ng/
L
(
3.4
×
10­
3
ppb)
(
ATSDR,
2000).

Tepper
et
al.
(
1995)
studied
contaminants
in
carpets
with
a
history
of
human­
health
related
complaints.
Pesticide
concentrations
in
the
carpets
were
determined
using
Soxhletextraction
(
with
6%
diethyl
ether/
hexane)
and
GC/
MS.
Trace
amounts
of
pesticides
were
detected
in
both
carpet
samples.
Aldrin
concentrations
extracted
from
the
first
carpet
ranged
from
ND­
83
:
g/
m2.
In
the
second
carpet,
extracts
contained
130
to
150
:
g/
m2
aldrin.
Estimates
of
aldrin
emissions
from
each
carpet
type
were
not
determined
in
this
study.

Dieldrin
Dieldrin
was
present
in
rainfall
measured
at
three
points
in
Canada
during
1984,
at
mean
concentrations
of
0.78
ng/
L
(
7.8
×
10­
4
ppb)
over
Lake
Superior,
0.27
ng/
L
in
New
Brunswick,
and
0.38
ng/
L
(
3.8
×
10­
4
ppb)
over
northern
Saskatchewan
(
Strachan,
1988).
Dieldrin
was
detected
in
rainfall
over
College
Station,
Texas,
at
average
concentrations
of
0.80
ng/
L
(
8
×
10­
4
ppb),
with
a
washout
ratio
(
concentration
in
rain/
concentration
in
air)
of
approximately
8.9
(
Atlas
and
Giam,
1988).

Dieldrin
concentrations
in
rainfall
were
collected
in
the
Great
Lakes
Basin
in
1986,
approximately
10
years
after
aldrin
and
dieldrin
use
was
restricted.
Dieldrin
was
detected
at
all
four
sites
and
in
more
than
60%
of
the
samples
at
mean
concentrations
ranging
from
0.41
to
1.81
ng/
L
(
4.1
×
10­
4
to
1.8
×
10­
3
ppb).
The
highest
concentrations
of
dieldrin
were
found
in
samples
collected
at
Pelee
Island
at
the
western
end
of
Lake
Erie,
with
a
maximum
concentration
of
5.9
ng/
L
(
5.9
×
10­
3
ppb)
(
ATSDR,
2000).

Tepper
et
al.
(
1995)
studied
contaminants
in
carpets
with
a
history
of
human­
health
related
complaints.
Pesticide
concentrations
in
the
carpets
were
determined
using
Soxhletextraction
(
with
6%
diethyl
ether/
hexane)
and
GC/
MS.
Trace
amounts
of
pesticides
were
detected
in
both
carpet
samples.
Dieldrin
concentrations
extracted
from
the
first
carpet
ranged
from
ND­
120
:
g/
m2.
In
the
second
carpet,
extracts
contained
190
to
230
:
g/
m2
dieldrin.
Dieldrin
emissions
from
each
carpet
type
were
not
determined
in
this
study.

Several
studies
have
reported
dieldrin
residues
in
sediments.
Composite
sediment
bed
samples
collected
from
24
navigation
pools
of
the
upper
Mississippi
River
in
1994
(
after
the
1993
flooding)
were
analyzed
for
organochlorine
pesticides.
While
dieldrin
was
detected
in
several
of
the
navigation
pools,
specific
concentrations
were
not
reported
(
Barber
and
Writer,
1998).

An
analysis
of
sediment
samples
taken
from
Lake
Ontario
in
1981
showed
that
dieldrin
levels
had
increased
from
approximately
0.026
mg/
kg
in
1970
to
0.048
mg/
kg
in
1980,
although
the
use
of
dieldrin
was
banned
in
much
of
the
Great
Lakes
Basin
in
the
early
1970s
(
Eisenreich
et
al.,
1989).
5­
15
Aldrin/
Dieldrin
 
February
2003
Eighty­
two
and
84
sediment
samples
were
collected
in
1994
and
1995,
respectively,
from
estuaries
along
the
Carolinian
Province
(
Cape
Henry,
Virginia,
to
St.
Lucie
Inlet,
Florida)
as
part
of
the
EPA
and
NOAA's
Environmental
Monitoring
and
Assessment
Program
(
EMAP).
Dieldrin
concentrations
ranged
from
0
to1.4
ng/
g
(
0
to
1.4
×
10­
2
mg/
kg)
in
1994,
and
from
0
to
38.5
ng/
g
(
0
to
3.9
×
10­
1
mg/
kg)
in
1995
(
Hyland
et
al.,
1998).

Bed
sediments
were
collected
from
16
sites
along
the
Lauritzen
Canal
and
Richmond
Harbor
of
the
San
Francisco
Bay
area
during
1994
to
study
the
distribution
of
contaminants
from
a
pesticide
processing
facility
point
source
(
also
a
National
Priorities
List
[
NPL]
site)
along
the
canal
into
the
San
Francisco
Bay.
Dieldrin
concentrations
in
sediments
(
up
to
5
cm
depths)
ranged
from
<
0.1
to
400
ng/
g
(
1.1
×
10­
4
to
0.4
mg/
kg)
dry
weight.
Concentrations
decreased
with
distance
from
the
head
of
the
canal,
as
the
three
sites
with
dieldrin
concentrations
above
11
ng/
g
(
1.1
×
10­
2
mg/
kg)
were
located
in
Lauritzen
Canal
(
Pereira
et
al.,
1996).

Burt
and
Ebell
(
1995)
analyzed
sediment
samples
from
an
industrial,
commercial,
and
recreational
area
off
the
coast
of
Perth,
Australia,
during
November
1991
for
organic
pollutants.
Dieldrin
was
detected
at
3
of
the
135
sites
sampled
at
concentrations
of
0.002
mg/
kg
dry
weight.

5.5
Summary
of
Exposure
to
Aldrin/
Dieldrin
in
Media
Other
Than
Water
Concentration
and
estimated
intake
values
for
aldrin
and
dieldrin
in
media
other
than
water
are
summarized
in
Tables
5­
3
to
5­
6
below.
Most
exposure
to
aldrin
and
dieldrin
for
the
general
population
and
agricultural
worker
subpopulation
appears
to
occur
through
diet.

Table
5­
3.
Summary
of
General
Population
Exposures
to
Aldrin
in
Media
Other
than
Water
Parameter
Medium
Food
Air
Soil
Adult
Child
Adult
Child
Adult
Child
Concentration
in
Medium
Non­
Fish
Food
(
NF):

0.0016
mg/
kg
Fish
and
Shellfish
(
F):

0.01
to
0.12
mg/
kg
4.5
x
10­
7
mg/
m3
NA1
Estimated
Daily
Intake
(
mg/
kg­
day)
NF:

3.0
x
10­
5
F:

2.9
x
10­
6
to
3.5
x
10­
5
NF:

1.3
x
10­
4
F:

4.0
x
10­
6
to
4.8
x
10­
5
1.3
x
10­
7
6.8
x
10­
7
­­
2
­­

1
NA
=
Not
Available.
2
­­
=
Unable
to
estimate
from
available
information.
5­
16
Aldrin/
Dieldrin
 
February
2003
Table
5­
4.
Summary
of
General
Population
Exposure
to
Dieldrin
in
Media
Other
than
Water
Parameter
Medium
Food
Air
Soil
Adult
Child
Adult
Child
Adult
Child
Concentration
in
Medium
Non­
Fish
Food
(
NF):

0.0015
mg/
kg
Fish
Food
(
F):

0.028
mg/
kg
1.6
x
10­
6
mg/
m3
0.03
to
1.1
mg/
kg
Estimated
Daily
Intake
(
mg/
kg­
day)
NF:

2.8
x
10­
5
F:

8.0
x
10­
6
NF:

1.3
x
10­
4
F:

1.1
x
10­
5
4.6
x
10­
7
2.4
x
10­
6
2.1
x
10­
8
to
7.9
x
10­
7
3.0
x
10­
7
to
1.1
x
10­
5
Table
5­
5.
Summary
of
Subpopulation
Exposures
to
Aldrin
in
Media
Other
than
Water
Parameter
Medium
Food
Air
Soil
Adult
Worker
Adult
Worker
Adult
Worker
Concentration
in
Medium
NA1
NA
NA
Estimated
Daily
Intake
(
mg/
kg­
day)
­­
2
­­
­­

1
NA
=
Not
Available.

2
­­
=
Unable
to
estimate
from
available
information.
5­
17
Aldrin/
Dieldrin
 
February
2003
Table
5­
6.
Summary
of
Subpopulation
Exposures
to
Dieldrin
in
Media
Other
than
Water
Parameter
Medium
Food
Air
Soil
Adult
Worker
Adult
Worker
Adult
Worker1
Concentration
in
Medium
0.015
to
0.028
mg/
kg
NA2
<
2
x10­
5
to
0.071
mg/
kg
high
end:
4.25
mg/
kg
Estimated
Daily
Intake
(
mg/
kg­
day)
2.8
x
10­
4
to
5.2
x
10­
4
­­
3
1.4
x
10­
10
to
2.9
x
10­
5
high
end:
2.9
x
10­
5
1
Estimates
are
intensive
contact
worker.

2
NA
=
Not
Available.

3
­­
=
Unable
to
estimate
from
available
information.
5­
18
Aldrin/
Dieldrin
 
February
2003
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6­
1
Aldrin/
Dieldrin
 
February
2003
6.0
TOXICOKINETICS
6.1
Absorption
Few
studies
pertaining
to
the
direct
measurement
of
the
absorption
of
aldrin
or
dieldrin
were
found
in
the
available
literature,
with
quantitative
human
data
being
especially
limited.
Dose­
related
increases
in
the
blood
and
adipose
tissue
levels
of
dieldrin
were
reported
for
volunteers
who
had
been
fed
approximately
0.0001,
0.0007,
or
0.003
mg/
kg­
day
of
dieldrin
for
18
to
24
months
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and
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Hunter
et
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1969).
After
18
months,
the
low,
intermediate,
and
high
exposures
resulted
in
blood
concentrations
of
dieldrin
that
had
increased
approximately
2­,
4­,
and
10­
fold
(
to
3,
5,
and
15
:
g/
L),
respectively.
The
authors
determined
that
under
steady­
state
conditions,
the
concentration
of
dieldrin
in
the
blood
(:
g/
L)
was
equal
to
approximately
8.6%
of
the
amount
ingested
(:
g/
day).
In
a
case
of
acute
poisoning,
one
of
two
children
who
ingested
dieldrin
died;
3
days
later,
the
blood
level
of
dieldrin
in
the
surviving
child
was
determined
to
be
0.27
ppm,
decreasing
to
0.11
ppm
within
2
weeks
(
Garrettson
and
Curley,
1969).
Concentrations
of
dieldrin
in
the
plasma
of
small
groups
of
pregnant
women
were
reported
to
range
from
0.0001
to
0.0061
ppm,
while
those
in
whole
cord
blood
of
newborns
ranged
from
0.0002
to
0.0015
ppm
(
Curley
and
Kimbrough,
1969;
Curley
et
al.,
1969).

Beyermann
and
Eckrich
(
1973)
conducted
inhalation
studies
with
aldrin
using
human
volunteers
that
suggested
approximately
50%
of
inhaled
aldrin
vapor
was
absorbed
and
retained
in
the
human
body.
However,
based
on
a
study
of
10
male
volunteers
who
were
exposed
to
measured
aldrin
vapor
concentrations
of
1.31
:
g/
m3,
followed
weeks
later
by
a
60­
minute
exposure
to
15.5
:
g/
m3,
actual
retention
may
have
been
closer
to
20%.
In
another
study,
apparently
healthy
workers
who
were
occupationally
exposed
to
aldrin
and
dieldrin
were
reported
to
have
a
mean
plasma
dieldrin
concentration
of
0.0185
ppm,
with
a
mean
of
5.67
ppm
stored
in
adipose
tissue
(
Hayes
and
Curley,
1968).
Although
uncertain,
exposure
was
likely
to
have
been
by
both
inhalation
and
dermal
contact.
Similarly,
in
a
study
discussed
more
fully
below,
Mick
et
al.
(
1971)
demonstrated
plasma
levels
of
aldrin
and
dieldrin
(
approximately
0.01
to
0.13
and
0.1
to
0.3
ppm,
respectively)
in
six
workers
who
had
formulated
2
million
lbs
of
aldrin
over
a
5­
week
period.
Stacey
and
Tatum
(
1985)
conducted
a
survey
study
of
women
in
pesticide­
treated
homes
that
demonstrated
a
correlation
between
home
treatment
and
dieldrin
levels
in
the
women's
breast
milk.

Many
distribution/
metabolism
studies
have
also
demonstrated
that
absorption
of
aldrin
and
dieldrin
occurs
in
animals
following
oral
exposure.
After
a
single
oral
dose
of
10
mg
aldrin/
kg
bw
was
given
to
neonatal
rats,
absorption
was
indicated
by
the
presence
of
aldrin
and/
or
dieldrin
in
various
tissues
over
the
succeeding
6
days
(
Farb
et
al.,
1973).
When
2
male
rats
were
given
4.3
:
g
of
radiolabeled
aldrin/
day
in
corn
oil
for
90
days
by
gavage,
3.6%
of
the
administered
total
dose
remained
in
the
carcass
24
hours
after
the
final
exposure
(
Ludwig
et
al.,
1964).
These
authors
estimated
that
approximately
10%
of
the
administered
dose
was
absorbed
by
the
gastrointestinal
(
GI)
tract.

Similarly,
Hayes
(
1974)
demonstrated
that
a
single
oral
dose
of
10
mg/
kg
bw
of
dieldrin
in
corn
oil
given
to
male
Sprague­
Dawley
rats
produced
consistent
concentrations
of
dieldrin
in
6­
2
Aldrin/
Dieldrin
 
February
2003
plasma
and
various
other
organs
and
tissues.
When
rats
were
fed
50
ppm
of
dieldrin
in
their
diet,
its
concentration
in
blood
and
liver
increased
for
the
first
9
days
before
then
remaining
fairly
constant
over
the
next
6
months.
Within
1
to
5
hours
after
orally
dosing
rats
with
radiolabeled
aldrin
or
dieldrin,
high
levels
of
radioactivity
were
detected
in
the
blood,
liver,
stomach,
and/
or
duodenum
(
Heath
and
Vandekar,
1964;
Iatropoulos
et
al.,
1975).
Heath
and
Vandekar
(
1964)
were
also
able
to
demonstrate
that
absorption
occurred
primarily
via
the
hepatic
portal
vein
and
not
the
thoracic
lymph
duct.

In
vivo
studies
on
the
inhalation
exposure
of
animals
to
either
aldrin
or
dieldrin
were
not
available,
but
Mehendale
and
El­
Bassiouni
(
1975)
demonstrated
that
aldrin
(
0.2
to
3.0
:
M)
was
taken
up
by
simple
diffusion
in
isolated,
perfused
rabbit
lungs.
Uptake
of
aldrin
was
biphasic,
a
slower
phase
following
the
initial
rapid
phase,
and
was
followed
by
a
slower
metabolism
to
dieldrin,
which
was
first
detected
3
minutes
after
initiation
of
the
experiment.

Several
studies
have
demonstrated
that
aldrin
and
dieldrin
can
be
absorbed
through
the
intact
skin
of
rabbits,
dogs,
monkeys,
and
humans
(
Shah
and
Guthrie,
1976;
Sundaram
et
al.,
1978a;
Fisher
et
al.,
1985;
ATSDR,
2000;
IPCS,
1989).
It
appears
to
occur
rapidly
in
humans,
with
aldrin
and
dieldrin
being
first
detected
in
the
urine
of
six
volunteers
just
4
hours
after
a
single
dermal
application
(
0.004
mg/
cm2)
of
the
radiolabeled
compounds
to
the
forearm
(
Feldmann
and
Maibach,
1974).
They
reported
that
approximately
8%
of
the
dermally
applied
compounds
(
in
acetone
vehicle)
were
absorbed
after
5
days.
The
accuracy
of
these
observations
has
been
questioned,
however,
as
the
dose
and
the
14C
recovery
in
the
urine
were
small,
the
major
route
of
excretion
was
the
feces
and
not
the
urine,
and
there
was
large
inter­
individual
variation.

In
female
rats,
aldrin
(
0.006,
0.06,
and
0.6
mg/
cm2)
was
rapidly
and
proportionally
absorbed
through
the
skin,
with
aldrin
and
dieldrin
detectable
in
the
skin
after
1
hour
at
all
three
dose
levels
(
Graham
et
al.,
1987).
In
vitro
exposure
of
rat
skin
strips
to
aldrin
showed
that
absorption
was
complete
after
80
minutes
(
Graham
et
al.,
1987).
In
rabbits,
dermal
absorption
was
demonstrated
from
fabric
that
had
been
impregnated
with
up
to
0.04%
dieldrin
(
Witherup
et
al.,
1961).

6.2
Distribution
As
a
result
of
its
relatively
rapid
conversion
to
dieldrin
(
see
Section
6.3),
aldrin
is
seldom
observed
in
human
tissues,
and
very
little
information
is
available
concerning
its
distribution
within
the
human
body
following
absorption
into
the
circulating
blood
(
ATSDR,
2000;
IPCS,
1989;
USEPA,
1992).
Given
their
hydrophobic
nature
and
high
solubilities
in
fat,
it
is
not
surprising
that
the
largest
concentrations
of
aldrin,
dieldrin,
and
their
metabolites
are
generally
found
in
adipose
tissue,
both
in
human
and
animal
studies
(
ATSDR,
2000;
IPCS,
1989;
USEPA,
1992,
1988,
1980).

Dale
and
Quinby
(
1963)
determined
the
concentrations
of
chlorinated
hydrocarbon
pesticides
in
the
body
fat
of
30
individuals­
28
from
the
general
population,
1
with
previous
aldrin
exposure
and
1
with
previous
DDT
exposure.
Mean
body
fat
dieldrin
concentration
(
±
SE)
for
the
general
population
was
0.15
±
0.02
µ
g/
g,
while
for
the
aldrin­
exposed
individual
it
was
6­
3
Aldrin/
Dieldrin
 
February
2003
0.36
µ
g/
g.
In
several
other
studies
of
the
same
era,
values
for
the
adipose
tissue
concentration
of
dieldrin
in
the
general
population
ranged
from
0.04
:
g/
g
(
India)
to
0.31
:
g/
g
(
U.
S.)
(
IARC,
1974b).
As
briefly
noted
in
the
section
on
absorption,
Hayes
and
Curley
(
1968)
examined
the
aldrin
and
dieldrin
concentrations
in
71
workers
involved
in
manufacturing
pesticides.
In
decreasing
order,
the
mean
concentrations
(
±
SE)
of
dieldrin
in
adipose
tissue,
urine
and
plasma
were
5.67
±
1.11,
0.0242
±
0.0063,
and
0.0185
±
0.0019
:
g/
g,
respectively;
these
were
significantly
higher
than
corresponding
values
reported
for
the
general
population.

In
a
study
conducted
on
male
volunteers,
3
men/
group
(
4
controls)
received
a
daily
oral
dose
of
either
0,
10,
50,
or
211
:
g
dieldrin
(
approximately
equivalent
to
0,
0.0001,
0.0007,
or
0.003
mg/
kg­
day)
for
18
months
(
Hunter
and
Robinson,
1967;
Hunter
et
al.,
1969).
The
50
and
211
:
g
groups
continued
to
receive
these
doses
for
another
6
months,
whereas
three
of
four
controls
and
the
10
:
g
group
were
switched
to
the
211
:
g
dose.
After
18
months,
concentrations
of
dieldrin
in
the
blood
of
the
low­,
intermediate­,
and
high­
dose
groups
had
increased
approximately
2­,
4­,
or
10­
fold
(
to
approximately
3,
5,
or
15
:
g/
L),
respectively.
It
was
noted
that
the
increase
in
the
low­
dose
group
had
essentially
been
achieved
by
5
months,
with
little
change
occurring
thereafter.
No
significant
increase
in
blood
dieldrin
concentration
during
the
18
to
24
month
period
was
noted
for
the
mid­
dose
group,
while
the
high­
dose
group
experienced
a
slight
increase
during
months
18
to
21,
but
nothing
significant
thereafter.
During
the
final
18­
to
24­
month
period,
the
control
and
low­
dose
subjects,
who
were
then
receiving
211
:
g
dieldrin/
day,
experienced
3­
fold
or
greater
increases
in
blood
concentrations
of
dieldrin.
After
18
months,
adipose
tissue
concentrations
of
dieldrin
in
the
low­,
intermediate­,
and
high­
dose
groups
had
increased
approximately
3­,
4­,
or
11­
fold
(
to
means
of
0.4,
0.7,
or
2
mg/
kg
tissue),
respectively.
An
apparent
further
increase
in
these
values
at
24
months
may
have
been
at
least
partly
related
to
sampling
techniques
(
IPCS,
1989).
Using
empirically
derived
relationships
between
the
amounts
of
dieldrin
ingested
and
those
found
in
the
blood
or
adipose
tissue,
the
authors
calculated
an
adipose
tissue
to
blood
distribution
ratio
under
steady
state
conditions
(
among
intake,
storage,
and
elimination)
of
136.

In
examining
tissue
samples
from
a
number
of
routine
autopsies,
De
Vlieger
et
al.
(
1968)
determined
the
mean
dieldrin
concentrations
in
adipose
tissue,
liver
tissue,
white
matter
of
the
brain,
and
gray
matter
of
the
brain
to
be
0.17,
0.03,
0.0061,
and
0.0047,
respectively.
Figure
6­
1,
taken
from
IPCS
(
1989),
represents
the
tentative
tissue
distribution
scheme
for
dieldrin
initially
proposed
by
De
Vlieger
et
al.
(
1968),
as
subsequently
recalculated
by
Jager
(
1970)
to
incorporate
the
empirical
formulas
of
Hunter
et
al.
(
1969).

Hunter
and
Robinson
(
1968)
demonstrated
that
the
leanest
subjects
had
both
the
highest
adipose
tissue
concentrations
of
dieldrin,
as
well
as
the
smallest
total
body
burdens;
however,
the
subjects
with
the
greatest
total
body
fat
retained
the
highest
proportion
of
the
total
exposure
dose
in
their
adipose
tissue.
As
no
increase
in
blood
levels
of
dieldrin
were
observed
during
surgical
stress
or
periods
of
complete
fasting,
these
authors
concluded
that
the
general
population
was
not
in
danger
of
intoxication
as
a
result
of
tissue
catabolism
during
periods
of
illness
or
weight
loss.
It
should
also
be
noted
that
when
Hunter
et
al.
(
1969)
followed
their
subjects
for
a
period
of
8
months
after
the
2­
year
exposure,
the
concentration
of
dieldrin
in
the
blood
was
observed
to
6­
4
Aldrin/
Dieldrin
 
February
2003
Figure
6­
1.
Distribution
Scheme
for
Dieldrin
Among
Blood
and
Various
Tissues
in
Humans
[
De
Vlieger
et
al.
(
1968)
as
Modified
by
Jager
(
1970);
From
IPCS
(
1989)]

decline
exponentially
with
an
approximate
half­
life
of
369
days.
There
were,
however,
significant
differences
among
individuals
in
the
rates
of
decline.
This
value
compares
with
a
mean
half­
life
of
266
days,
which
was
estimated
for
dieldrin
in
the
blood
of
15
occupationally
exposed
workers
during
a
3­
year
period
following
termination
of
their
exposure
(
Jager,
1970).
In
the
Garettson
and
Curley
(
1969)
study
of
aldrin
poisoning
in
children
that
was
noted
in
Section
6.2,
47
ppm
dieldrin
was
measured
in
a
fat
specimen
taken
3
days
after
the
exposure;
6
months
later
this
value
had
declined
to
15
ppm,
where
it
remained
after
8
months.

A
study
of
women
and
their
offspring
during
labor
demonstrated
that
placental
transfer
of
dieldrin
can
occur
(
Polishuk
at
al.,
1977).
Higher
concentrations
of
dieldrin
were
observed
in
fetal
blood
(
1.22
mg/
kg)
than
in
maternal
blood
(
0.53
mg/
kg),
and
in
the
placenta
(
0.8
mg/
kg)
than
in
the
uterus
(
0.54
mg/
kg).

In
the
previously
discussed
(
Section
6.2)
study
of
six
workers
occupationally
exposed
for
5
weeks
via
inhalation
and
dermal
contact
to
aldrin,
Mick
et
al.
(
1971)
examined
the
distribution
of
aldrin
and
dieldrin
among
erythrocytes,
plasma,
and
the
alpha­
and
beta­
lipoprotein
fractions
of
blood.
The
epoxidation
of
aldrin
to
dieldrin
led
to
higher
plasma
concentrations
of
dieldrin
(
approximately
0.1
to
0.3
ppm)
than
aldrin
(
approximately
0.01
to
0.13
ppm).
Average
dieldrin
residues
were
approximately
four
times
higher
in
plasma
than
in
erythrocytes
and
this
ratio
tended
to
increase
with
increasing
concentrations
of
dieldrin
in
the
blood.
Typically,
higher
dieldrin
levels
were
associated
with
the
beta­
lipoprotein
fraction
than
with
the
alpha­
lipoprotein
fraction.
The
in
vitro
study
of
human
blood
fractions
by
Skalsky
and
Guthrie
(
1978)
also
demonstrated
that
dieldrin
could
bind
to
albumin
and
beta­
lipoprotein.
6­
5
Aldrin/
Dieldrin
 
February
2003
Distribution
of
aldrin
and
dieldrin
has
been
studied
in
a
number
of
animal
species
(
ATSDR,
2000;
IARC,
1974a,
b;
IPCS,
1989;
USEPA,
1992,
1988,
1980).
Exposure
of
mammals
to
aldrin
leads
to
deposition
of
dieldrin
in
their
adipose
tissue
(
Jager,
1970).
Deichmann
et
al.
(
1975)
fed
Swiss­
Webster
mice
diets
containing
0,
5,
or
10
ppm
aldrin
(
approximately
equivalent
to
0,
0.75,
and
1.5
mg/
kg
bw,
based
on
Leyman
[
1959])
over
the
course
of
7
generations
(
from
weaning
to
age
260
days
for
each
generation,
except
F
4;
see
below).
After
4
generations
of
aldrin
feeding,
metabolic
conversion
to
dieldrin
and
subsequent
retention
led
to
significantly
increased
levels
of
dieldrin
in
abdominal
fat
and
carcass
total
lipids.
Significantly
increased
retention
of
dieldrin
in
the
whole
carcass
was
observed
for
the
F
1
generation,
with
smaller
and
not
statistically
significant
increases
observed
for
the
F
2
and
F
3
generations.
Dieldrin
concentration
in
F
0
carcass
total
lipids
was
60
mg/
kg,
whereas
the
F
1
+
F
2
+
F
3
grouped
means
for
males
and
females
were
100
and
132
mg/
kg,
respectively.
Female
mice
thus
retained
higher
residue
levels
in
their
body
fat
than
male
mice.
From
weaning
through
day
260,
the
F
4
generation
was
fed
only
the
aldrin­
free
control
diet,
and
the
pesticide
residues
that
it
absorbed
in
utero
and
through
lactation
were
found
to
have
been
completely
excreted
by
the
time
of
sacrifice.
Dieldrin
concentrations
in
F
5
pups
were
<
1
mg/
kg;
aldrin­
containing
diets
were
resumed
upon
the
weaning
of
these
pups,
with
the
findings
from
the
F
4
through
F
6
generations
largely
paralleling
those
from
the
F
0
through
F
2
generations.

Two
male
Wistar
rats
were
given
daily
doses
of
4.3
:
g
14C­
aldrin
by
gavage
for
3
months,
and
then
sacrificed
24
hours
after
the
final
dose
(
Ludwig
et
al.,
1964).
Relative
to
the
total
administered
amount
of
radiolabel,
the
amounts
recovered
in
the
carcass,
abdominal
fat,
and
other
tissues
were
3.60,
1.77,
and
1.83%,
respectively.
A
steady
state
among
intake,
storage,
and
excretion
was
reportedly
achieved
after
53
days.
Ratios
of
dieldrin
to
aldrin
found
in
the
carcass
and
the
abdominal
fat
were
approximately
15:
1
and
18:
1,
respectively.
In
neonatal
Sprague­
Dawley
rats
given
a
single
dose
of
10
mg
aldrin/
kg
bw,
aldrin
was
detectable
up
to
6
days
later
in
the
stomach
and
small
intestine,
but
only
for
3
days
in
the
kidneys
(
Farb
et
al.,
1973).
Aldrin
concentrations
in
the
liver
increased
during
the
first
6
hours
to
a
maximum
of
13%
of
the
administered
dose,
then
declined
to
<
0.1%
by
72
hours.
The
only
metabolite
identified
in
the
liver
was
dieldrin,
which
was
detectable
as
early
as
2
hours
post­
treatment
and
which
reached
a
maximum
31%
of
the
administered
radiolabeled
dose
after
24
hours.

Deichmann
et
al.
(
1969)
administered
0.6
mg
aldrin/
kg
bw/
day
in
corn
oil
to
6
male
beagle
dogs
for
10
months.
Dieldrin
concentrations
in
body
fat
and
the
liver
were
observed
to
progressively
increase
to
70
and
20
ppm,
respectively,
and
then
decline
over
the
12
months
postexposure
to
25
and
6
ppm,
respectively.
In
a
related
study,
aldrin
was
administered
by
capsule
to
3
male
beagles
(
0.3
mg/
kg
bw)
and
4
female
beagles
(
0.15
or
0.3
mg/
kg
bw),
5
days/
week
for
14
months
(
Deichmann
et
al.,
1969,
1971).
During
the
last
10
months
of
exposure,
dieldrin
concentrations
in
the
blood
and
subcutaneous
fat
for
the
high­
dose
animals
were
0.042
to
0.183
and
37
to
208
mg/
L,
respectively;
those
for
the
low­
dose
females
were
0.040
to
0.130
and
12
to
67
mg/
kg,
respectively.
The
apparent
subcutaneous
fat
to
blood
partition
ratio
was
thus
approximately
1000.

An
extensive
comparative
study
of
the
distribution
and
metabolism
of
dieldrin
and
its
metabolites
in
male
CFE
rats
and
male
CF
1
and
LACG
mice
was
conducted
by
Hutson
(
1976).
6­
6
Aldrin/
Dieldrin
 
February
2003
14C­
dieldrin
was
administered
as
a
single
oral
dose
to
animals,
either
with
or
without
a
4­
week
pretreatment
of
dieldrin
(
20
mg/
kg
diet
for
rats,
10
mg/
kg
diet
for
mice),
and
the
animals
were
sacrificed
8
days
later.
Concentrations
of
dieldrin
were
much
higher
in
the
fat
than
in
the
liver
or
kidneys
of
all
animals,
and
were
higher
in
the
fat
and
liver
of
mice
(
11.6
and
0.94
mg/
kg)
than
of
rats
(
5.6
and
0.11
mg/
kg).
Tissue
levels
of
a
number
of
dieldrin
metabolites
(
see
Section
6.3)
were
also
assessed,
including
the
6,7­
dihydroxy
(
diol)
derivative
that
was
found
to
be
below
the
level
of
detection
(<
0.02
mg/
kg)
in
the
fat,
liver,
and
kidneys
of
all
animals.
Concentrations
of
the
9­
hydroxy
metabolite
were
very
low
(<
0.03
mg/
kg)
in
the
fat
and
kidneys,
but
small
amounts
were
found
in
the
livers
of
both
mouse
strains.
The
pentachloroketone
metabolite
was
found
in
rat
liver
in
small
amounts
and
in
much
larger
amounts
in
the
kidneys
of
rats,
with
or
without
pretreatment;
small
concentrations
were
also
found
in
the
fat
of
both
groups.
In
both
strains
of
mice,
this
metabolite
was
undetectable
or
present
in
only
very
small
amounts
in
the
fat,
liver,
and
kidneys
in
the
absence
of
pretreatment;
with
pretreatment,
higher
concentrations
were
observed
(
e.
g.,
~
1.3
mg/
kg
in
fat).

At
1
to
2
hours
after
dosing
rats
with
radiolabeled
dieldrin,
Heath
and
Vandekar
(
1964)
observed
the
highest
concentration
of
dieldrin
in
adipose
tissue;
high
levels
were
also
seen
in
the
liver
and
kidneys,
with
moderate
concentrations
found
in
the
brain.
It
was
also
recoverable
from
the
stomach,
small
and
large
intestines,
and
the
feces
after
1
hour.
Following
dietary
exposure
to
radiolabeled
dieldrin
for
8
hours,
high
levels
of
radioactivity
were
detected
in
the
kidneys
of
treated
rats
(
Matthews
et
al.,
1971).
While
somewhat
more
radioactivity
was
found
in
the
kidneys,
lungs,
stomachs,
and
intestines
of
males,
in
general,
for
the
other
organs
and
tissues,
females
had
3
to
4
times
the
radioactivity
as
did
males.
Similar
results
were
observed
in
a
9­
week
(
5
day/
week)
feeding
study
with
Osborne­
Mendel
rats
(
Dailey
et
al.,
1970).
Adipose
tissue
was
again
shown
to
be
the
principal
storage
depot
for
dieldrin,
with
significant
levels
also
found
in
the
kidneys,
liver,
lungs,
and
adrenals;
lowest
levels
were
seen
in
the
spleen,
brain,
and
heart.
With
the
exception
of
the
kidneys,
more
radioactivity
was
retained
in
the
tissues
of
females
than
males.
In
a
single
oral
dose
rat
study
by
Iatropoulos
et
al.
(
1975),
radiolabeled
dieldrin
was
rapidly
taken
up
by
the
liver
during
the
first
3
hours,
then
redistributed
in
a
biphasic
manner
to
adipose
tissue
(
the
majority),
kidneys,
lymph
nodes,
etc.
The
lymphatic
system
appeared
to
be
the
principal
redistribution
pathway
and
parallel
dieldrin
increases
in
the
lymph
nodes
and
adipose
tissue
suggested
an
equilibrium
between
lymph
and
depot
fat.

Female
Osborne­
Mendel
rats
were
fed
a
diet
containing
technical
grade
dieldrin
(
87%
purity)
at
a
concentration
of
50
mg/
kg
diet
(
approximately
2.5
mg/
kg
bw/
day)
for
6
months
(
Deichmann
et
al.,
1968).
Rats
were
sacrificed
at
various
times
up
to
183
days
and
the
retention
of
dieldrin
in
blood,
liver,
and
fat
was
examined.
Tissue
levels
increased
rapidly
over
the
first
9
days
in
the
blood
and
liver,
and
over
the
first
16
days
in
fat;
thereafter,
concentrations
fluctuated
some
but
did
not
appear
to
significantly
increase
further.
Over
the
final
4
months,
distribution
ratios
and
mean
concentrations
were:
blood
=
1
(
0.240
mg/
L),
liver
=
28
(
6.8
mg/
kg),
and
fat
=
665
(
159.5
mg/
kg).

Groups
of
Carworth
Farm
E
rats
(
25/
sex;
45
controls/
sex)
were
fed
0,
0.1,
1.0,
or
10
mg
dieldrin/
kg
diet
for
2
years
(
Walker
et
al.,
1969).
Animals
were
sacrificed
after
26,
52,
78,
and
104
weeks,
and
tissue
levels
of
dieldrin
in
the
blood,
brain,
liver,
and
fat
were
determined.
6­
7
Aldrin/
Dieldrin
 
February
2003
Approximate
plateau
levels
were
reached
by
week
26;
tissue
uptake
ratios
(
tissue
concentration/
diet
concentration)
of
dieldrin
for
the
3
female
exposure
groups
were
0.056
(
blood),
0.19
(
brain),
0.35
(
liver),
and
8.8
(
fat),
and
were
significantly
higher
than
the
corresponding
values
for
males.
Estimated
partition
ratios
(
tissue
concentration/
blood
concentration)
for
male/
female
animals
were
1/
1
(
blood),
3.3/
2.6
(
brain),
7.8/
5.9
(
liver),
and
104/
137
(
fat).
After
104
weeks,
tissue
levels
were
found
to
be
generally
2
to
10
times
higher
in
females
than
in
males
(
Table
6­
1).
Robinson
et
al.
(
1969)
fed
Carworth
rats
10
mg
dieldrin/
kg
diet
for
8
weeks,
then
a
control,
dieldrin­
free
diet
for
up
to
an
additional
12
weeks.
Again,
the
concentration
of
dieldrin
was
found
to
be
substantially
the
greatest
in
adipose
tissue,
followed
in
descending
order
by
that
in
the
liver,
brain,
and
blood.
Following
exposure,
the
decline
of
dieldrin
concentrations
in
the
tissues
was
approximately
exponential,
with
half­
lives
in
adipose
tissue
and
brain
of
10.3
and
3
days,
respectively.
Elimination
from
the
liver
occurred
in
a
rapid
and
then
a
slower
phase,
with
respective
half­
lives
of
1.3
and
10.2
days;
similar
values
were
estimated
for
the
blood.

Three
groups
of
Sprague­
Dawley
rats
(
2/
sex)
were
fed
diets
containing
0.04
mg
14Cdieldrin
kg
plus
0,
0.16
or
1.96
mg/
kg
of
unlabeled
dieldrin
(
totals
of
0.04,
0.2,
or
2.0
mg
dieldrin/
kg
diet)
for
39
weeks
(
Davison,
1973).
For
all
three
groups
upon
sacrifice,
whole
carcass
radioactivity
as
a
percentage
of
administered
dose
was
significantly
higher
in
females
than
males
(
means
of
6.9
versus
2.1%,
respectively).

Table
6­
1.
Distribution
of
Dieldrin
in
Rats
after
104
Weeks1
Sex
Dieldrin
Concentration
(
mg/
kg)

Diet
Blood2
Fat2
Liver2
Brain2
Male
0.0
0.0009
0.0598
0.0059
0.0020
0.1
0.0021
0.2594
0.0159
0.0069
1.0
0.0312
1.493
0.1552
0.1040
10.0
0.1472
19.72
1.476
0.4319
Female
0.0
0.0015
0.3112
0.0112
0.0077
0.1
0.0065
0.8974
0.0348
0.0224
1.0
0.0861
13.90
0.4295
0.2891
10.0
0.3954
57.81
2.965
1.130
1
Walker
et
al.
(
1969);
as
modified
from
USEPA
(
1980).
2
Geometric
mean
values.
6­
8
Aldrin/
Dieldrin
 
February
2003
Baron
and
Walton
(
1971)
fed
male
Osborne­
Mendel
rats
25
mg
dieldrin/
kg
diet
(
approximately
1.25
mg/
kg
bw)
for
8
weeks.
They
reported
that
an
equilibrium
level
of
50
mg
dieldrin/
kg
had
been
achieved
in
adipose
tissue
by
week
8,
which
upon
return
to
a
dieldrin­
free
diet,
rapidly
declined
with
an
estimated
half­
life
of
4
to
5
days.
Within
15
days
after
the
cessation
of
exposure
to
75
ppm
dieldrin
in
the
diet,
levels
in
the
adipose
tissue
of
rats
had
fallen
to
half
that
seen
after
12
months
of
exposure
(
Robinson
and
Roberts,
1968).
In
a
study
by
Hayes
(
1974),
male
Sprague­
Dawley
rats
received
a
single
dose
of
10
mg/
kg
bw
of
technical
dieldrin
(
86%
purity).
At
intervals
up
to
240
hours
post­
dosing,
animals
were
sacrificed
and
tissue
levels
of
dieldrin
were
determined.
In
plasma,
dieldrin
concentrations
reached
a
maximum
of
~
0.5
mg/
L
after
2
hours,
fluctuated
from
0.2
to
0.5
mg/
L
up
to
48
hours,
then
declined
to
~
0.01
mg/
L
at
240
hours.
Maximum
levels
were
reached
in
the
brain
after
4
hours
(~
1
mg/
kg),
remaining
more
or
less
constant
through
48
hours,
then
declining
to
a
low
level
(<
0.2
mg/
kg)
by
240
hours;
similar
time
courses
were
reported
for
muscle,
kidneys,
and
the
liver.
A
slower
rise
of
dieldrin
concentration
was
observed
in
retroperitoneal
fat,
with
4
and
24
hours
values
being
~
10
and
40
mg/
kg,
respectively;
after
48
hours,
a
decline
similar
to
those
for
plasma
and
the
brain
was
observed.
For
the
4­
and
16­
hour
data,
Hayes
(
1974)
set
the
dieldrin
concentrations
in
the
brain
equal
to
1.00,
then
calculated
the
relative
concentrations
for
the
other
tissues
evaluated
(
Table
6­
2).

Table
6­
2.
Relative
Tissue
Levels
of
Dieldrin
in
the
Rat
Following
a
Single
Oral
Dose1
Hour
Brain
Muscle
Liver
Kidney
Plasma
Fat
4
1.00
±
0
0.62
±
0.05
2.30
±
0.11
1.55
±
0.22
0.20
±
0.02
7.20
±
1.18
16
1.00
±
0
0.55
±
0.06
3.17
±
0.25
2.02
±
0.56
1.35
±
1.11
17.96
±
3.23
1
Dieldrin
dose
of
10
mg/
kg,
in
corn
oil
(
Hayes,
1974;
as
modified
from
USEPA,
1980).

In
vitro
studies
using
rats
and
rabbits
have
reportedly
examined
the
partitioning
of
14Cdieldrin
related
radioactivity
among
the
soluble
protein
and
cellular
components
of
the
blood
(
IPCS,
1989).
Radioactivity
was
principally
found
in
erythrocytes
(
associated
with
hemoglobin
and
an
unknown
constituent)
and
the
plasma,
with
much
lower
levels
found
in
leukocytes,
platelets,
and
erythrocyte
membranes.
In
rat
serum,
it
electrophoresed
with
pre­
and
postalbumin
whereas
in
rabbit
serum,
it
was
associated
with
albumin
and
"­
globulin.
Ichinose
and
Kurihara
(
1985)
demonstrated
in
vitro
that
transport
of
dieldrin
between
rat
hepatocytes
and
the
extracellular
medium
occurs
much
more
rapidly
than
does
intra­
hepatocyte
metabolic
transformation.

Several
studies
have
been
conducted
on
the
distribution
of
dieldrin
in
dogs
(
Richardson
et
al.,
1967;
Keane
and
Zavon,
1969;
Walker
et
al.,
1969).
In
three
beagles
fed
0.1
mg
dieldrin/
kg
bw
for
128
days,
blood
levels
of
dieldrin
increased
curvilinearly
to
an
approximate
plateau
of
about
0.130
mg/
L
by
day
93
(
Richardson
et
al.,
1967).
One
week
post
exposure,
measured
tissue
6­
9
Aldrin/
Dieldrin
 
February
2003
levels
of
dieldrin
were
0.150
mg/
L
(
blood),
1.090
mg/
kg
(
heart),
4.420
mg/
kg
(
liver),
2.330
mg/
kg
(
kidneys),
14.030
mg/
kg
(
pancreas),
0.710
mg/
kg
(
spleen),
1.227
mg/
kg
(
lungs),
25.333
mg/
kg
(
fat),
and
0.566
mg/
kg
(
muscle).
There
was
reported
a
highly
significant
linear
correlation
between
the
logarithms
of
exposure
duration
and
blood
dieldrin
level.
Keane
and
Zavon
(
1969)
orally
dosed
4
male
and
2
female
mongrel
dogs
with
1
mg
dieldrin/
kg
bw
(
in
corn
oil)
for
5
days,
then
with
0.2
mg/
kg
bw
for
the
next
54
days.
Small
but
significant
increases
in
dieldrin
concentration
were
observed
in
the
blood
of
all
animals
for
days
7
to
59
(
samples
taken
twice
weekly
from
day
7
onward).
Subcutaneous
fat
biopsies
were
taken
on
days
16
and
50
and
the
fat
to
blood
partition
ratios
were
216
and
117,
respectively.

In
addition
to
the
rat
study
discussed
previously,
Walker
et
al.
(
1969)
orally
dosed
beagle
dogs
(
5/
sex)
by
gel
capsule
with
0,
0.005,
or
0.05
mg
dieldrin/
kg
bw
(
equivalent
to
0,
0.1,
or
1.0
ppm
in
the
diet)
for
2
years.
Blood
dieldrin
levels
increased
during
the
first
12
to
18
weeks,
reaching
a
plateau
during
weeks
18
to
76.
Thereafter,
significant
deviations
from
this
apparent
asymptotic
value
were
observed;
while
the
reasons
for
this
were
not
clear,
a
tendency
toward
higher
dieldrin
concentrations
was
also
noted
in
the
control
animals.
Uptake
and
partition
ratios
(
previously
defined)
for
males
were
0.06
and
1.0
(
blood),
0.22
and
3.7
(
brain),
4.4
and
10
(
liver),
and
10.0
and
169
(
adipose
tissue).
In
contrast
to
the
rat
study,
no
significant
sex
differences
in
uptake
were
apparent.

Mueller
et
al.
(
1975a)
administered
14C­
dieldrin
(
2.5
mg/
kg
bw)
to
two
female
rhesus
monkeys
via
intravenous
injection,
and
to
two
males
via
a
single
oral
dose
of
either
0.36
or
0.5
mg/
kg
bw.
Females
and
males
were
sacrificed
at
75
and
10
days
post
exposure,
respectively.
In
all
animals,
the
highest
radioactivity
was
observed
in
the
adipose
tissue,
bone
marrow
and
liver,
with
only
a
relatively
small
amount
present
in
the
brain
(~
2%
that
of
adipose
tissue).
Metabolites,
though
present
in
the
bile,
were
not
detected
in
the
organs
or
tissues
examined.
In
another
primate
study,
groups
of
male
rhesus
monkeys
were
fed
diets
containing
0,
0.01,
0.1,
0.5,
or
1.0
ppm
of
technical
grade
dieldrin
for
70
to
74
months
(
Wright
et
al.,
1978).
Several
monkeys
were
started
at
a
5.0
ppm
dose,
but
when
1
died
at
4
months,
the
others
were
reduced
to
a
dose
of
2.5
ppm
for
the
next
5
months,
then
1.75
ppm
for
a
further
64
months.
In
one
animal
from
this
group,
the
1.75
ppm
dose
was
gradually
increased
back
up
to
5.0
ppm
at
month
23,
where
it
remained
for
another
46
months.
This
study
focused
on
interactions
of
dieldrin
with
the
liver,
and
mean
concentrations
of
dieldrin
in
the
livers
of
the
various
groups
ranged
from
1.2
mg/
kg
(
the
0.01
ppm
group)
to
23.3
mg/
kg
(
the
single
5,
2.5,
1.75ÿ5.0
ppm
monkey).
When
the
distribution
of
dieldrin
in
the
liver's
various
subcellular
fractions
was
examined,
~
60%
was
localized
in
the
microsomal
fraction,
~
12.5%
in
the
soluble
fraction,
and
~
9%
each
in
the
nuclear,
mitochondrial,
and
lysosomal
fractions.
It
was
noted
that
at
dietary
intakes
of
about
0.1
ppm,
tissue
concentrations
of
dieldrin
in
rhesus
monkeys
and
humans
were
similar.
However,
when
compared
to
male
rats,
liver
concentrations
of
dieldrin
were
200
times
higher
in
these
monkeys
at
a
dose
only
twice
as
high,
suggesting
a
relatively
slow
metabolic
clearance
and
a
relatively
high
liver
tolerance
to
dieldrin
in
this
primate
species.

With
respect
to
dermal
exposure,
most
of
the
dieldrin
that
is
absorbed
through
the
skin
of
guinea
pigs,
dogs,
and
monkeys
has
been
found
to
accumulate
in
adipose
tissue
(
Sundaram
at
al.,
1978a,
b).
In
guinea
pigs
dermally
exposed
for
6
months
to
concentrations
of
0.0001
to
0.1%,
the
6­
10
Aldrin/
Dieldrin
 
February
2003
highest
tissue
levels
were
observed
in
adipose
tissue,
with
lesser
concentrations
appearing
in
the
liver
and
brain
(
Sundaram
et
al.,
1978b).
After
52
weeks
of
exposure
to
fabric
strips
containing
up
to
0.04%
dieldrin,
rabbits
also
evidenced
a
slight
accumulation
of
the
compound
in
omental
and
renal
fat
(
Witherup
et
al.,
1961).

Distribution
of
dieldrin
residues
among
the
blood,
brain,
liver,
and
subcutaneous
fat
in
rats
following
intraperitoneal
injection
was
not
found
to
be
significantly
different
from
that
seen
after
oral
exposure,
i.
e.,
the
highest
levels
were
again
observed
in
adipose
tissue
(
Lay
et
al.,
1982).
Transplacental
transport
of
dieldrin
has
been
reported
to
occur
to
a
significant
extent
in
mice
following
intramuscular
injection
(
Baeckstroem
et
al.,
1965)
and
after
intravenous
injections
in
rats
(
Eliason
and
Posner,
1971)
and
rabbits
(
Hathway
et
al.,
1967).
In
pregnant
mice
exposed
intramuscularly
to
14C­
dieldrin,
the
highest
radioactivities
were
observed
in
the
adipose
tissue,
liver,
intestines,
and
mammary
glands,
while
moderate
activities
were
reported
for
the
ovaries
and
brain
(
Baeckstroem
et
al.,
1965).
Transfer
across
the
placenta
was
indicated
by
the
moderate
levels
that
were
also
found
in
fetal
liver,
fat,
and
intestines.
Finally,
numerous
studies
suggest
that
the
toxicokinetics
of
aldrin
and
dieldrin
in
most
domesticated
animals
are
at
least
broadly
similar
to
those
seen
in
laboratory
species
(
IPCS,
1989).

Although
apparently
not
a
major
transformation
product
in
mammals,
photodieldrin
is
likely
a
significant
photodegradation
product
and
microbial
metabolite
of
dieldrin
in
the
environment
(
see
Chapter
3),
and
therefore
several
studies
have
examined
its
distribution
pattern
in
mammals.
Collectively,
subacute
and
subchronic
studies
in
rats
(
Dailey
et
al.,
1970;
Walker
et
al.,
1971;
Walton
et
al.,
1971)
and
mice
(
Brown
et
al.,
1967)
have
demonstrated
that
females
accumulate
2­
to
15­
fold
higher
concentrations
of
photodieldrin
than
do
males
in
adipose
and
other
tissues
with
the
exception
of
the
kidneys.
The
estimated
half­
life
of
photodieldrin
in
adipose
tissue
is
also
longer
in
female
rats
(
2.6
days)
than
in
male
rats
(
1.7
days)
(
Brown
et
al.,
1967).
When
a
single
oral
dose
of
photodieldrin
was
administered
to
one
male
and
one
female
dog,
tissue
levels
were
again
reported
to
be
significantly
higher
in
the
female
than
in
the
male,
with
the
exception
of
in
the
liver
(
Brown
et
al.,
1967).
In
contrast,
a
3­
month
feeding
study
in
dogs
demonstrated
dose­
related
concentrations
in
the
liver,
adipose
tissue,
and
kidneys
that
were
similar
in
both
males
and
females
(
Walker
et
al,
1971);
additionally,
the
kidney
levels
of
photodieldrin
and
pentachloroketone
(
PCK;
see
Figures
6­
2
and
6­
3,
Table
6­
3,
and
associated
text
in
Section
6.3)
were
approximately
0.1
to
0.2
mg/
kg,
or
about
1
to
3
orders
of
magnitude
lower
than
those
observed
in
rats.

6.3
Metabolism
Radomski
and
Davidow
(
1953)
first
reported
the
epoxidation
of
aldrin
to
dieldrin.
Since
that
time,
many
studies
in
a
substantial
number
of
organisms
have
shown
this
to
be
the
initial
and
principal
step
in
the
biotransformation
of
aldrin;
the
reaction
is
mediated
by
mixed­
function
oxidases,
sometimes
referred
to
as
aldrin­
epoxidase,
that
are
known
to
be
found
in
substantial
quantities
in
the
endoplasmic
reticulum
of
hepatocytes
in
vertebrates
(
ATSDR,
2000;
IPCS,
1989;
USEPA,
1992,
1988).
Perhaps
understandably,
no
real
metabolism
studies
of
aldrin
or
dieldrin
in
humans
were
located
or
available,
so
data
on
the
human
metabolism
of
these
compounds
is
sparse.
Excretion
data
in
humans
have
provided
some
insight,
however,
as
the
6­
11
Aldrin/
Dieldrin
 
February
2003
9­
hydroxydieldrin
metabolite
was
detected
in
the
feces
of
workers
having
occupational
exposure
to
aldrin
and
dieldrin
(
Richardson
and
Robinson,
1971).
Some
additional
excretion
data
from
humans
on
these
compounds
and
their
metabolites
are
presented
in
Section
6.4.

A
variety
of
metabolites
have
been
isolated
from
microorganisms,
invertebrates,
and
vertebrates,
and
the
three­
dimensional
chemical
structures
of
many
of
these
are
presented
in
Figure
6­
2.
Their
trivial,
or
common,
names
are
listed
in
the
companion
Table
6­
3.
Those
metabolic
transformations
thought
to
be
most
important
in
laboratory
animals
are
illustrated
in
Figure
6­
3,
which
again
provides
three­
dimensional
chemical
structures,
as
well
as
some
of
the
enzymes
implicated
in
these
pathways.
A
discussion
of
some
of
the
more
important
underlying
animal
and
in
vitro
studies
follows.

Winteringham
and
Barnes
(
1955)
first
demonstrated
the
epoxidation
of
aldrin
to
dieldrin
(
Figure
6­
2,
compounds
I
and
II;
Figure
6­
3)
in
mice,
and
were
able
to
show
that
this
conversion
occurred
more
rapidly
in
males
than
in
females;
while
other
metabolites
were
not
observed,
methodological
limitations
may
have
hindered
the
detection
of
polar
compounds.
The
formation
of
dieldrin
from
aldrin
was
also
noted
early
on
in
cattle,
pigs,
sheep,
rats,
and
poultry
(
Bann
et
al.,
1956),
and
Soto
and
Deichmann
(
1967)
reported
that
subsequent
to
the
intravascular
administration
of
aldrin
to
dogs,
approximately
30%
was
converted
to
dieldrin
during
the
first
24
hours
post­
exposure.
Using
rabbit
lung
perfusates,
Mehendale
and
El­
Bassiouni
(
1975)
were
able
to
demonstrate
dose­
dependent,
in
vitro
metabolism
of
aldrin
to
dieldrin
within
the
endoplasmic
reticulum;
at
low
doses,
up
to
70%
conversion
occurred
during
the
first
hour.
Following
dermal
application
of
0.1
to
10
mg/
kg
to
rats,
the
skin
has
also
been
shown
capable
of
converting
aldrin
to
dieldrin
(
Graham
et
al.,
1987).
Dieldrin
was
detected
in
the
skin
as
soon
as
1
hour
after
application,
and
enzyme
saturation
was
suggested
because
the
highest
percentage
of
conversion
occurred
at
the
lowest
dose.
The
authors
estimated
that
up
to
10%
conversion
to
dieldrin
by
skin
enzymes
could
result
in
rats
from
the
percutaneous
absorption
of
aldrin.
Graham
et
al.
(
1987)
were
also
able
to
demonstrate
the
in
vitro
dermal
conversion
of
aldrin
to
dieldrin
in
studies
employing
mouse
skin
microsomal
preparations
and
whole­
skin
strips
from
rats.

Using
liver
microsome
preparations
from
male
and
female
rats,
Wong
and
Terriere
(
1965)
were
able
to
demonstrate
the
conversion
of
aldrin
to
dieldrin
via
nicotine
adenine
dinucleotide
phosphate
(
NADPH)­
dependent,
heat­
labile
mixed
function
oxidases,
that
the
reaction
proceeded
more
rapidly
in
the
microsomes
from
male
rat
livers
than
in
those
from
females,
and
that
it
could
be
inhibited
by
pesticide
synergists,
such
as
sesamex.
These
observations
were
largely
confirmed
by
Nakatsugawa
et
al.
(
1965)
using
microsome
preparations
from
male
rats
and
rabbits;
they
also
reported
that
dieldrin
did
not
undergo
further
microsomal
metabolism,
that
epoxidase
activity
in
liver
preparations
was
10­
fold
higher
than
in
lung
preparations,
and
that
no
such
activity
was
observed
in
preparations
from
the
kidney,
spleen,
pancreas,
heart,
or
brain.
Wolff
et
al.
(
1979)
demonstrated
a
three­
fold
increase
in
dieldrin
formation
with
microsomes
taken
from
phenobarbital­
treated
rats,
whereas
amounts
were
substantially
decreased
in
microsomal
preparations
made
from
rats
pretreated
with
3­
methylcholanthrene.
These
results
suggested
that
aldrin
epoxidation
involved
cytochrome
P­
450
rather
than
cytochrome
P­
448.
6­
12
Aldrin/
Dieldrin
 
February
2003
Kurihara
et
al.
(
1984)
have
demonstrated
that
cultures
of
rat
hepatocytes
are
effective
in
carrying
out
the
epoxidation
of
aldrin
to
dieldrin.
In
other
in
vitro
studies,
Lang
et
al.
(
1986)
investigated
the
epoxidation
of
aldrin
to
dieldrin
in
hepatic
and
various
extra­
hepatic
tissues
in
the
rat.
Unlike
the
liver,
many
organs
and
tissues
contain
little
cytochrome
P­
450
activity,
prompting
these
authors
to
look
for
the
presence
of
an
alternative
oxidative
pathway
mediated
by
prostaglandin
endoperoxide
synthase
(
PES)
in
liver,
lung,
seminal
vesicle,
and
subcutaneous
granulation
tissues.
In
a
two­
step
process,
PES
utilizes
cyclooxygenase
activity
to
catalyze
the
bisdioxygenation
of
arachidonic
acid
to
prostaglandin
G
2
(
PGG
2),
which
is
subsequently
reduced
to
prostaglandin
H
2
(
PGH
2)
via
hydroperoxidase
activity;
it
is
during
this
latter
step
that
xenobiotics
(
e.
g.,
aldrin)
may
be
co­
oxidized
(
i.
e.,
epoxidized).
In
hepatocytes
and
liver
microsomes,
aldrin
epoxidation
was
reported
to
be
completely
NADPH­
dependent,
whereas
in
lung
microsomes,
two
pathways
appeared
involved
(
Lang
et
al.,
1986).
The
NADPH­
dependent
and
arachidonic
acid­
dependent
aldrin
epoxidation
activities
were
1.5
and
0.3%,
respectively,
of
the
activities
observed
in
liver
preparations.
Aldrin
epoxidation
was
stimulated
by
arachidonic
acid
and
inhibited
by
the
cyclooxygenase­
specific
inhibitor
indomethacin,
in
microsomal
preparations
from
seminal
vesicle
and
subcutaneous
granulation
tissues.
Therefore,
the
PES
pathway
would
appear
to
be
an
alternative
route
for
aldrin
epoxidation
in
extra­
hepatic
tissues.

In
some
early
work
with
rabbits,
Korte
(
1963)
was
able
to
identify
one
of
the
metabolites
of
aldrin
as
aldrin
trans­
diol
(
Figure
6­
2,
compound
IV;
Table
6­
3,
Figure
6­
3).
Heath
and
Vandekar
(
1964)
reported
that
the
principal
route
of
excretion
in
rats
was
the
feces,
that
little
dieldrin
was
excreted
unchanged,
and
that
a
somewhat
polar
metabolite
could
be
found
in
the
feces,
along
with
other
polar
metabolites
in
both
the
feces
and
urine.
After
feeding
14C­
aldrin
to
rats
for
3
months,
Ludwig
et
al.
(
1964)
found
aldrin,
dieldrin,
and
unidentified
hydrophilic
metabolites
in
the
urine;
these
latter
constituted
75
and
95%
of
the
radioactivity
excreted
in
the
urine
and
feces,
respectively.
Two
different
metabolites
were
detected
in
the
feces,
with
one
of
them
and
a
third
metabolite
also
detected
in
the
urine.
In
rabbits
dosed
orally
with
14C­
dieldrin
for
21
weeks,
Korte
and
Arent
(
1965)
isolated
6
urinary
metabolites,
the
major
one
(
86%)
being
identified
as
6,7­
trans­
dihydroxydihydroaldrin,
or
the
aforementioned
aldrin
trans­
diol.
This
enzymatic
product
of
epoxide
hydrase,
however,
appears
to
be
of
relatively
minor
importance
in
most
other
species
(
IPCS,
1989).

Other
than
in
mice
and
rabbits,
aldrin
trans­
diol
has
reportedly
been
found
in
rhesus
monkeys
and
chimpanzees
(
Mueller
et
al.,
1975b),
and
its
glucuronide
conjugation
product
was
detected
in
liver
microsomal
preparations
from
rats
or
rabbits
incubated
in
the
presence
of
14Cdieldrin
and
uridine
diphosphoglucuronic
acid
(
UDPGA)
(
Matthews
and
Matsumura,
1969).
This
water
soluble
metabolite
accounted
for
approximately
45%
of
the
total
radioactivity,
while
the
unconjugated
form
was
also
found
to
be
present
in
vitro.
Matthews
and
Matsumura
(
1969)
had
additionally
fed
male
rats
a
diet
containing
dieldrin
for
a
month,
and
had
noted
a
minor
metabolite
present
in
both
the
feces
and
the
urine.
Comparative
thin
layer
chromatography
in
conjunction
with
the
in
vitro
results
indicated
this
compound
to
be
the
aldrin
trans­
diol
in
the
conjugated
and/
or
unconjugated
forms.
6­
13
Aldrin/
Dieldrin
 
February
2003
Figure
6­
2.
Metabolites
of
Aldrin
and
Dieldrin
(
from
IPCS,
1989).
For
the
Identity
(
Trivial
Chemical
Names)
of
These
Compounds,
See
Table
6­
3
6­
14
Aldrin/
Dieldrin
 
February
2003
Table
6­
3.
Trivial
Chemical
Names
of
Aldrin,
Dieldrin
and
Their
Metabolites
(
as
Identified
in
Figure
6­
2)
1
ID
Code
(
Fig.
6­
2)
Chemical
Structure
Trivial
Names
Alternative
1
Alternative
2
I
Aldrin
HHDN
II
Dieldrin
HEOD
III
Photodieldrin
IV
Aldrin
trans­
diol
6,7­
trans­
dihydroxydihydroaldrin
V
Aldrin
dicarboxylic
acid
VI
9­
Hydroxy
dieldrin
9­
Hydroxydieldrin
VII
(
Bridged)
Pentachloroketone
PCK
(
or
Klein's
metabolite)

VIII
Dechloro­
aldrin
dicarboxylic
acid
IX
Dieldrin
ketone
X
Photodieldrin
ketone
XI
Photodieldrin
trans­
diol
Caged
aldrin
trans­
diol
XII
Photoaldrin
dicarboxylic
acid
Caged
aldrin
acid
XIII
Photoaldrin
1
Taken
principally
from
IPCS
(
1989)
and
ATSDR
(
2000).
6­
15
Aldrin/
Dieldrin
 
February
2003
Figure
6­
3.
Proposed
Principal
Metabolic
Pathways
for
Aldrin
and
Dieldrin
(
from
ATSDR,
1993,
as
Adapted
from
USEPA,
1987)

The
conjugation
of
aldrin
trans­
diol
with
glucuronic
acid
and/
or
its
further
oxidation
to
aldrin
dicarboxylic
acid
(
Figure
6­
2,
compound
XII;
Table
6­
3,
Figure
6­
3)
have
also
been
reported
by
Baldwin
et
al.
(
1972),
Hutson
(
1976),
and
Oda
and
Mueller
(
1972).
Formation
of
the
cis­
diol
and
its
epimerization
to
the
trans­
diol
have
been
demonstrated
to
occur
in
rat
microsomes
(
McKinney
et
al.,
1973).

In
two
studies
involving
the
feeding
of
dieldrin
to
male
rats
for
7
months
(
Richardson
et
al.,
1968)
or
1
month
(
Matthews
and
Matsumura,
1969),
two
major
metabolites
were
isolated
from
the
urine
and
feces.
The
fecal
metabolite
proved
to
be
9­
hydroxy
dieldrin
(
Figure
6­
2,
compound
VI;
Table
6­
3,
Figure
6­
3);
this
reaction
was
found
to
be
catalyzed
by
liver
microsomal
monooxygenases
in
rats,
and
to
be
inhibited
by
the
monooxygenase
inhibitor,
sesamex
(
Matthews
and
Matsumura,
1969).
With
the
exception
of
the
rabbit,
in
most
of
the
6­
16
Aldrin/
Dieldrin
 
February
2003
species
studied
(
i.
e.,
mice,
rats,
sheep,
rhesus
monkeys,
chimpanzees),
this
has
been
the
principal
metabolite
that
has
been
found
(
Feil
et
al.,
1970;
Mueller
et
al.,
1975b).
It
has
been
detected
in
the
feces,
and
either
free
or
conjugated
in
the
urine.
After
dosing
sheep
with
14C­
dieldrin,
Hedde
et
al.
(
1970)
isolated
six
hexane­
soluble
and
two
water­
soluble
urinary
metabolites,
postulating
that
one
of
the
latter
was
the
glucuronide
conjugate
of
aldrin
trans­
diol
(
Figure
6­
3).
Two
of
the
hexane­
soluble
metabolites
were
subsequently
identified
as
aldrin
trans­
diol
and
9­
hydroxy
dieldrin
(
Feil
et
al.,
1970).
The
glucuronide
conjugate
of
9­
hydroxy
dieldrin
is
formed
both
in
vivo
and
in
vitro
and
has
been
isolated
in
the
bile
of
rats
(
Chipman
and
Walker,
1979);
passing
through
the
bile
duct
into
the
lower
intestines,
it
is
largely
converted
there
into
the
free
9­
hydroxy
metabolite
before
being
excreted
in
the
feces
(
Hutson,
1976).
When
dieldrin
is
incubated
in
vitro
with
rat
liver
microsomes
in
the
presence
of
UDPGA,
9­
hydroxy
dieldrin
glucuronide
is
reported
to
form
rapidly
via
the
consecutive
actions
of
microsomal
monooxygenase
and
uridine
diphosphoglucuronyl
transferase
(
Hutson,
1976;
Matthews
et
al.,
1971).
As
evidence
of
species
differences
in
the
rates
of
metabolism
of
dieldrin,
a
higher
ratio
of
9­
hydroxy
14C­
dieldrin
to
14C­
dieldrin
has
been
observed
in
rats
than
in
mice,
indicative
of
a
more
rapid
hydroxylation
reaction
in
the
former
(
Hutson,
1976).

The
second
major
metabolite
(
i.
e.,
the
one
found
in
the
urine)
that
was
reported
in
the
rat
studies
of
Richardson
et
al.
(
1968)
and
Matthews
and
Matsumura
(
1969)
has
been
identified
as
pentachloroketone,
or
PCK
(
Figure
6­
2,
compound
VII;
Table
6­
3,
Figure
6­
3).
Also
known
as
Klein's
metabolite,
it
has
been
found
mainly
in
the
urine
and
kidneys
of
male
rats,
but
only
in
small
amounts
in
female
rats,
mice,
and
other
species
(
Baldwin
et
al.,
1972;
Damico
et
al.,
1968;
Hutson,
1976;
Klein
et
al.,
1968;
Matthews
et
al.,
1971;
Richardson
et
al.,
1968).
Male
rats
have
been
found
to
metabolize
dieldrin
3
to
4
times
more
rapidly
than
females
(
Matthews
et
al.,
1971),
a
difference
that
has
been
ascribed
to
males'
greater
ability
to
convert
dieldrin
to
its
more
polar
metabolites,
including
9­
hydroxy
dieldrin
(
ATSDR,
2000)
and
especially
PCK
(
USEPA,
1980).

Comparative
metabolism
studies
on
male
CFE
rats
and
male
CF
1
and/
or
LACG
mice
revealed
that
much
greater
quantities
of
the
PCK
derivative
were
produced
in
the
rat
than
in
either
mouse
strain,
smaller
amounts
of
polar
urinary
metabolites
were
produced
in
the
mice,
and
aldrin
trans­
diol
was
found
in
the
feces
and
a
dicarboxylic
acid
derivative
in
the
urine
of
all
animals;
both
rats
and
mice
produced
9­
hydroxy
dieldrin
(
Baldwin
et
al.,
1972;
Hutson,
1976).
In
their
study
of
rats
dosed
with
radiolabeled
dieldrin,
Matthews
et
al.
(
1971)
found
that
the
greatest
percentage
of
radioactivity
in
the
feces
of
both
males
and
females
came
from
9­
hydroxy
dieldrin,
with
aldrin
trans­
diol
and
a
second,
unidentified
polar
metabolite
also
present.
Significant
amounts
of
PCK
was
found
in
the
urine
of
male
rats,
with
initially
some
aldrin
transdiol
and
unchanged
dieldrin.
In
female
rats,
most
of
the
activity
in
urine
was
associated
with
aldrin
trans­
diol
and,
initially,
up
to
20%
with
dieldrin.

It
should
also
be
noted
that
when
photodieldrin
(
Figure
6­
2,
compound
III;
Table
6­
3),
itself
a
degradation
product
of
dieldrin,
was
fed
to
rats
for
13
weeks,
it
and
PCK
were
isolated
from
blood,
brain,
liver,
and
adipose
tissue
(
Baldwin
and
Robinson,
1969).
When
administered
orally
or
intraperitoneally
5
days/
week
for
12
weeks,
PCK
and
small
amounts
of
other
more
polar
metabolites
were
found
in
the
urine
of
rats
(
Klein
et
al.,
1970).
In
a
female
rhesus
monkey
given
daily
oral
doses
of
14C­
photodieldrin
for
175
days,
photodieldrin
trans­
diol
(
Figure
6­
2,
6­
17
Aldrin/
Dieldrin
 
February
2003
compound
XI;
Table
6­
3)
and
its
glucuronide
conjugate
were
identified
in
the
urine,
and
possibly
only
the
diol
in
the
feces
(
Nohynek
et
al.,
1979).
A
third
metabolite,
found
both
in
the
urine
and
feces,
was
speculated
to
be
a
mono­
hydroxy
derivative
of
photodieldrin.

Finally,
oral
administration
of
dieldrin
has
been
shown
capable
of
inducing
hepatic
mixed
function
oxidases
(
Kohli
et
al.,
1977).
Baldwin
et
al.
(
1972)
have
also
been
able
to
demonstrate
some
induction
in
the
CFE
male
rat
(
but
not
in
the
CF
1
male
mouse)
by
prefeeding
low
doses
of
dieldrin
for
3
weeks.
It
is
relevant
to
keep
this
observation
in
mind
when,
for
example,
comparing
the
results
of
long­
term
versus
acute
animal
studies,
or
considering
the
potential
effects
of
aldrin
or
dieldrin
exposure
in
humans
who
are
chronically
exposed
to
at
least
low
doses
of
mixed
function
oxidase
inducers
(
USEPA,
1980).

6.4
Excretion
Much
of
the
available
information
on
the
excretion
of
aldrin
and
dieldrin
has
already
been
introduced
in
the
previous
sections
describing
their
absorption,
distribution,
and
metabolism.
Some
of
this
information
will
again
be
briefly
mentioned
in
this
section,
along
with
additional
detail
in
some
cases
and
supplemented
with
a
number
of
additional
studies.
These
compounds
have
been
found
in
general
to
be
excreted
primarily
in
the
feces,
but
also
to
some
extent
in
the
urine,
in
the
form
of
metabolites
that
are
more
polar
than
the
parent
compounds.
When
exposure
is
kept
constant,
equilibrium
levels
of
aldrin,
dieldrin,
and
their
metabolites
are
generally
achieved
in
most
organs
and
tissues.
Body
burdens
will
fluctuate
in
accordance
with
increases
and
decreases
in
exposure
concentration.

Although
the
data
is
naturally
much
less
extensive
than
in
animals,
excretion
in
humans
following
exposure
to
aldrin
or
dieldrin
appears
to
occur
largely
through
the
bile
and
feces.
In
an
early
study
of
occupationally
exposed
workers,
Cueto
and
Hayes
(
1962)
were
able
to
detect
the
presence
of
dieldrin
and
some
of
its
metabolites
in
their
urine.
Somewhat
later,
Cueto
and
Biros
(
1967)
reported
that
the
mean
concentrations
of
dieldrin
found
in
the
urine
of
five
men
and
five
women
from
the
general
population
were
0.8
±
0.2
and
1.3
±
0.1mg/
L,
respectively.
These
concentrations
were
compared
with
those
of
male
workers
(
a
total
of
14)
who
were
deemed
to
have
either
low
(
5
males),
medium
(
4
males),
or
high
(
5
males)
occupational
exposure
to
dieldrin
and
other
chlorinated
insecticides.
The
respective
urinary
concentrations
of
the
three
worker
groups
were
5.3,
13.8,
and
51.4
mg/
L.
In
another
study
of
workers
occupationally
exposed
to
various
chlorinated
pesticides,
the
concentrations
of
aldrin
detected
in
14
urine
samples
were
all
less
than
0.2
mg/
L,
while
those
for
dieldrin
ranged
from
1.3
to
66.0
mg/
L
(
Hayes
and
Curley,
1968).
Two
reports
have
described
detecting
9­
hydroxy
dieldrin
in
the
feces
of
seven
workers
having
occupational
exposure
to
aldrin
and
dieldrin
(
Richardson,
1971;
Richardson
and
Robinson,
1971).
The
mean
and
range
of
fecal
9­
hydroxy
dieldrin
concentrations
measured
in
the
seven
workers
were
1.74
and
0.95
to
2.80
mg/
kg,
respectively,
while
those
determined
from
five
males
of
the
general
population
were
0.058
and
0.033
to
0.12
mg/
kg,
respectively.
Dieldrin
at
a
mean
concentration
of
0.18
mg/
kg
was
also
detected
in
the
feces
of
the
workmen,
but
it
could
not
be
detected
in
samples
from
the
general
population.
Examination
of
the
urine
for
dieldrin
and
four
known
metabolites
led
the
authors
to
conclude
that
urinary
excretion
was
a
6­
18
Aldrin/
Dieldrin
 
February
2003
minor
pathway
in
human
males,
although
they
failed
to
examine
the
urine
for
glucuronide
or
other
conjugates
of
the
potential
hydroxy
metabolites.

In
a
study
by
Hunter
et
al.
(
1969),
12
human
volunteers
ingested
various
amounts
of
dieldrin
for
up
to
24
months;
dieldrin
concentrations
in
blood
and
adipose
tissue
were
monitored
during
this
exposure
period,
as
were
the
blood
concentrations
for
an
additional
8
months.
For
3
of
the
volunteers,
blood
dieldrin
concentrations
reportedly
did
not
change
significantly;
for
the
remaining
9,
the
mean
half­
life
of
dieldrin
in
the
blood
was
estimated
to
be
369
days
(
a
range
of
141
to
592
days).
Though
determined
with
a
limited
number
of
samples,
this
estimate
was
far
longer
than
the
value
of
less
than
10
days
that
had
been
reported
in
animal
studies.
In
an
unpublished
study
by
DeJonge,
it
was
reported
(
Jager,
1970)
that
in
workers
who
had
had
previously
high
exposures
to
aldrin/
dieldrin,
and
thus
high
concentrations
of
the
compounds
in
their
blood
before
being
transferred
to
other
areas,
the
mean
half­
life
of
dieldrin
in
the
blood
had
been
calculated
to
be
0.73
years
(
or
approximately
266
days).
This
estimate
was
reportedly
based
on
measurements
taken
every
6
months
for
3
years
following
cessation
of
exposure.
It
agrees
reasonably
well
with
that
of
Hunter
et
al.
(
1969),
which
was
derived
using
limited
data.

Feldman
and
Maibach
(
1974)
demonstrated
that
7.7%
of
a
dose
of
14C­
dieldrin
(
4
:
g/
cm2
in
acetone),
applied
once
to
the
arm
of
volunteers,
was
excreted
in
the
urine
over
a
5­
day
period;
similarly,
3.3%
of
a
single
intravenous
injection
was
excreted
in
the
urine
over
the
same
period.
Finally,
dieldrin
can
be
excreted
via
lactation
in
nursing
mothers,
and
concentrations
ranging
from
1
to
29
ppb
have
been
reported
in
human
milk
samples
taken
from
women
in
various
countries
around
the
globe
(
Curley
and
Kimbrough,
1969;
Schecter
et
al.,
1989;
IARC,
1974b).

In
one
of
the
early
animal
studies
examining
the
metabolism
and
excretion
of
these
compounds,
Ludwig
et
al.
(
1964)
gave
male
Wistar
rats
daily
oral
doses
of
14C­
aldrin
(
4.3
:
g,
or
about
0.2
mg/
kg
diet)
for
up
to
3
months.
They
reported
that
approximately
9
times
as
much
radioactivity
was
excreted
in
the
feces
as
in
the
urine
(
urinary
excretion
increasing
from
~
2%
during
week
1
to
9
to
10%
during
week
12).
As
a
percent
of
administered
daily
dose,
excretion
increased
from
31%
on
day
2,
to
about
80%
during
week
2,
to
100%
by
weeks
8
to
12,
indicating
that
a
steady
state,
saturation
level
had
been
reached
in
the
animals.
Once
exposure
was
discontinued,
excretion
of
radiolabeled
compounds
diminished
rapidly;
24
hours,
6
weeks,
and
12
weeks
after
the
final
dose,
88,
98,
and
>
98%
of
the
total
administered
dose
had
been
excreted.
Urine
and
fecal
extracts
were
examined
by
paper
chromatography,
which
indicated
that
aldrin
content
in
both
urine
and
feces
decreased
during
the
exposure
period
and
afterward,
while
that
of
dieldrin
somewhat
increased.
Hydrophilic
metabolites
increased
during
exposure,
constituting
after
12
weeks
about
75
and
95%
of
the
radioactivity
excreted
in
feces
and
urine,
respectively.
Contrary
to
the
predominance
of
fecal
excretion
seen
in
this
rat
study,
it
has
been
reported
that
male
rabbits
administered
14C­
aldrin
excreted
more
radioactivity
in
their
urine
than
in
their
feces
(
IARC,
1974a).
In
rabbits
orally
administered
14C­
dieldrin
for
21
weeks,
Korte
and
Arent
(
1965)
observed
that
right
after
the
exposure
period
(
week
22),
42%
of
the
total
administered
radioactivity
had
been
excreted,
with
2
to
3
times
as
much
via
the
urine
as
the
feces.

In
female
rats
infused
for
2.5
to
5
hours
with
total
doses
of
8
to
16
mg/
kg
bw
of
36Cldieldrin
approximately
70
and
10%
of
the
administered
doses
were
recovered
over
the
ensuing
6­
19
Aldrin/
Dieldrin
 
February
2003
42
days
in
the
feces
and
the
urine,
respectively,
indicating
that
the
predominant
route
of
excretion
was
via
the
bile
(
Heath
and
Vandekar,
1964).
The
authors
also
noted
that
dietary
restriction
markedly
increased
the
blood
dieldrin
concentration
as
fat
stores
were
mobilized.
Comparable
findings
were
observed
in
male
rats
with/
without
biliary
fistulas
that
received
single
intravenous
doses
of
14C­
dieldrin
(
0.25
mg/
kg
bw)
(
Cole
et
al.,
1970).
After
7
days,
about
80%
of
the
administered
dieldrin
dose
had
been
excreted
in
the
feces.
At
1,4,
and
7
days
post­
exposure
in
the
rats
with
biliary
fistulas,
approximately
30,
60,
and
>
90%,
respectively,
of
the
administered
dose
had
been
excreted
via
the
bile.
In
experiments
with
isolated
perfused
rat
livers,
about
20%
of
the
perfused
dieldrin
dose
was
collected
in
the
bile
during
an
8­
hour
period
(
Cole
et
al.,
1970),
and
the
rate
of
biliary
excretion
in
those
isolated
from
males
was
found
to
be
approximately
three
times
greater
than
in
those
from
females
(
Klevay,
1970).
Chipman
and
Walker
(
1979)
reported
that
in
rats
receiving
dieldrin
intraperitoneally,
pretreatment
with
phenobarbital
increased
the
rate
of
biliary
excretion.

Dailey
et
al.
(
1970)
reported
that
following
exposure
to
radiolabeled
dieldrin,
excretion
of
radioactivity
via
urine
and
feces
was
higher
in
male
rats
than
in
female
rats,
a
finding
confirmed
in
a
39­
week
study
by
Davison
(
1973).
The
latter
study
also
indicated
that
the
maximal
excretion
of
radioactivity
occurred
during
the
6th
week
of
exposure,
regardless
of
the
dieldrin
dose,
and
that
a
steady
state
condition
existed
from
weeks
6
through
39.
Matthews
et
al.
(
1971)
found
a
10­
fold
higher
level
of
radioactivity
in
kidneys
isolated
from
males
than
from
females
in
rats
that
had
been
fed
14C­
dieldrin.
In
male
kidneys,
most
of
the
radioactivity
was
associated
with
PCK,
whereas
in
female
kidneys
only
dieldrin
was
detected.
This
greater
ability
of
male
rats
to
convert
dieldrin
to
its
more
polar
metabolites,
especially
PCK,
was
thought
to
underly
the
three­
to
fourfold
more
rapid
metabolism
of
dieldrin
that
is
observed
in
male
versus
female
rats.

Following
the
single
oral
administration
of
0.5
mg
14C­
dieldrin/
kg
bw
to
mice,
rats,
rabbits,
rhesus
monkeys,
and
one
chimpanzee,
urine
and
fecal
samples
were
collected
for
10
days
and
analyzed
(
Mueller
et
al.,
1975b).
They
reported
the
main
route
of
excretion
to
be
the
feces
for
all
species
except
the
rabbit,
accounting
for
95,
95,
~
18,
79,
and
79%
of
the
amounts
excreted,
respectively.
The
ratios
of
fecal
to
urinary
excretion
are
thus
approximately
19:
1
for
rats
and
mice,
1:
5
for
rabbits,
and
4:
1
for
rhesus
monkeys
and
the
chimpanzee.
Ten
days
after
dosing,
the
total
amounts
of
radioactivity
excreted
were
37%
(
mice),
11%
(
rats),
2%
(
rabbits),
20%
(
rhesus
monkeys),
and
6%
(
the
chimpanzee)
of
the
total
administered
dose.
In
all
five
species,
the
principal
metabolites
were
9­
hydroxy
dieldrin
and
aldrin
trans­
diol;
unchanged
dieldrin,
9­
hydroxy
dieldrin,
and
its
glucuronide
were
reported
to
predominate
in
rats,
rhesus
monkeys,
and
the
chimpanzee,
whereas
mice
and
rabbits
displayed
higher
amounts
of
aldrin
trans­
diol.
The
glucuronide
conjugate
of
aldrin
trans­
diol
was
identified
in
the
urine
of
the
rabbits
and
rhesus
monkeys,
and
aldrin
dicarboxylic
acid
was
noted
as
a
minor
metabolite
in
the
feces
of
rats,
rhesus
monkeys,
and
the
chimpanzee.
As
noted
previously,
Klein
et
al.
(
1968)
had
also
detected
PCK
in
the
urine
of
rats
fed
1.25
mg
aldrin/
kg/
day.

Baldwin
et
al.
(
1972)
compared
the
excretion
of
dieldrin
in
the
CF
1
mouse
and
the
CFE
rat
and
found
the
amounts
of
labeled
dieldrin
excreted
after
7
to
8
days
were
similar.
The
feces
contained
about
10
times
the
radioactivity
found
in
the
urine,
and
50
to
70%
of
the
administered
dose
was
excreted
during
the
1­
week
collection
period.
As
noted
previously,
the
proportion
of
6­
20
Aldrin/
Dieldrin
 
February
2003
various
metabolites
varied
between
the
two
species,
a
principal
difference
being
that
PCK
was
found
in
significant
amounts
in
rat
urine,
but
was
not
detected
in
mouse
urine.
Hutson
(
1976)
conducted
a
similar
study
on
male
CFE
rats
and
male
CF
1
and
LACG
mice
after
a
single
oral
dose,
with
or
without
a
period
of
dieldrin
pretreatment
(
see
Section
6.3).
Dieldrin
pretreatment
modestly
increased
the
percentage
fecal
excretion
(
of
the
total
administered
radiolabeled
dose)
from
62.4
to
69%
in
the
rats,
had
no
effect
in
the
LACG
mice
(
51.5%),
and
substantially
increased
it
in
CF
1
mice
from
27.2
to
48.8%.
Urinary
excretion
in
the
rats
was
5.5
to
6.6%,
whereas
it
was
much
lower
in
the
mice
(
0.42
to
2.6%).
In
the
male
rats
and
CF
1
mice,
the
amount
of
urinary
aldrin
dicarboxylic
acid
was
low
compared
with
that
of
PCK
+
dieldrin,
while
in
LACD
mice
it
was
twice
as
high.
A
much
higher
proportion
of
an
unidentified
metabolite
was
excreted
in
the
urine
of
both
mouse
strains
than
in
that
of
the
rat.
In
rats,
the
major
fecal
metabolite
was
9­
hydroxy
dieldrin
with
or
without
pretreatment;
in
both
strains
of
mice,
however,
it
became
a
major
fecal
metabolite
only
after
pretreatment.

In
a
study
of
sheep
dosed
with
14C­
dieldrin,
excretion
of
radioactivity
was
higher
in
the
feces
than
in
the
urine
(
Hedde
et
al.,
1970).
These
authors
noted
that
in
two
very
fat
sheep,
the
ratio
of
labeled
dieldrin
in
feces
to
that
in
urine
was
>
10:
1,
but
in
two
thin
sheep
receiving
the
same
dose,
it
was
only
slightly
greater
than
1:
1.
Only
0.25%
of
the
total
dose
was
exhaled
as
14CO
2,
and
after
5
to
6
days
of
collection,
less
than
50%
of
the
administered
dose
was
recovered.

For
more
information
on
the
relatively
rapid
loss
of
dieldrin
and/
or
its
metabolites
from
various
organs
and
tissues,
refer
to
the
relevant
studies
previously
discussed
in
Section
6.2
(
e.
g.,
Robinson
et
al.,
1969;
Barron
and
Walton,
1971).
Finally,
the
excretion
of
photodieldrin
has
been
explored
in
rats
(
Dailey
et
al.,
1970)
and
monkeys
(
Nohynek
et
al.,
1979).
After
12
weeks
of
daily
dosing
with
14C­
photodieldrin
in
the
rat,
urinary
excretion
was
found
to
be
significantly
higher
in
males
than
in
females,
and
to
gradually
increase
during
the
12­
week
exposure
period.
Fecal
excretion
was
initially
lower
in
females,
but
became
greater
during
the
latter
part
of
the
study
(
Dailey
et
al.,
1970).
After
orally
dosing
rhesus
monkeys
for
70
to
76
days
with
radiolabeled
photodieldrin,
a
steady
state
between
intake
and
excretion
was
reported
(
Nohynek
et
al.,
1979).
At
the
end
of
exposure,
the
animals
had
excreted
about
50%
of
the
cumulative
dose,
and
an
additional
30%
was
excreted
during
the
next
100
days.
During
dosing,
photodieldrin
was
a
major
fecal
metabolite,
and
20
to
50%
of
the
radioactivity
was
excreted
in
the
urine;
this
amount
increased
to
60%
when
the
dosing
ceased.
When
one
male
and
one
female
rhesus
monkey
were
given
a
single
intravenous
injection
of
radiolabeled
photodieldrin,
excretion
remained
high
during
the
first
7
days,
but
then
rapidly
decreased.
By
day
21,
approximately
45
and
34%
of
the
administered
dose
had
been
excreted
in
the
male
and
female,
respectively.
6­
21
Aldrin/
Dieldrin
 
February
2003
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in
USEPA,
1992).

Stacey,
C.
I.
and
T.
Tatum.
1985.
House
treatment
with
organochlorine
pesticides
and
their
level
in
milk
!
Perth,
Western
Australia.
Bull.
Environ.
Contam.
Toxicol.
35:
202­
208
(
as
cited
in
ATSDR,
2000).

Sundaram,
K.
S.,
V.
N.
Damodaran,
and
T.
A.
Venkitasubramanian.
1978a.
Absorption
of
dieldrin
through
monkey
and
dog
skin.
Indian
J.
Exp.
Biol.
16:
101­
103
(
as
cited
in
ATSDR,
2000;
IPCS,
1989).

Sundaram,
K.
S.,
V.
N.
Damodaran,
and
T.
A.
Venkitasubramanian.
1978b.
Absorption
of
dieldrin
through
skin.
Indian
J.
Exp.
Biol.
16:
1004­
1007
(
as
cited
in
ATSDR,
2000).
6­
28
Aldrin/
Dieldrin
 
February
2003
TOXLINE.
2000a.
Abstracts
of:
Deichmann,
W.
B.,
W.
E.
MacDonald
and
D.
A.
Cubit.
1975.
Dieldrin
and
DDT
in
the
tissues
of
mice
fed
aldrin
and
DDT
for
seven
generations.
Arch.
Toxicol.
34(
3):
173­
182.
Retrieved
Oct.
4,
2000.
Bethesda,
MD:
National
Library
of
Medicine,
Specialized
Information
Services
Division,
Toxicology
and
Environmental
Health
Information
Program,
TOXLINE
database.

TOXLINE.
2000b.
Abstract
of:
Hunter,
C.
G.,
J.
Robinson
and
M.
Roberts.
1969.
Pharmoacodynamics
of
dieldrin
(
HEOD).
Ingestion
by
human
subjects
for
18
to
24
months,
and
postexposure
for
eight
months.
Arch.
Environ.
Health
18(
1):
12­
21.
Retrieved
Oct.
4,
2000.
Bethesda,
MD:
National
Library
of
Medicine,
Specialized
Information
Services
Division,
Toxicology
and
Environmental
Health
Information
Program,
TOXLINE
database.

TOXLINE.
2000c.
Abstracts
of:
Wright,
A.
S.,
C.
Donninger,
R.
D.
Greenland,
K.
L.
Stemmer
and
M.
R.
Zavon.
1978.
The
effects
of
prolonged
ingestion
of
dieldrin
on
the
livers
of
male
rhesus
monkeys.
Ecotoxicol.
Environ.
Safety
1(
4):
477­
502.
Retrieved
Oct.
2,
2000.
Bethesda,
MD:
National
Library
of
Medicine,
Specialized
Information
Services
Division,
Toxicology
and
Environmental
Health
Information
Program,
TOXLINE
database.

USEPA.
1992.
U.
S.
Environmental
Protection
Agency.
Aldrin
drinking
water
health
advisory.
Washington,
DC:
USEPA
Office
of
Water.

USEPA.
1988.
U.
S.
Environmental
Protection
Agency.
Dieldrin
health
advisory.
Washington,
DC:
USEPA
Office
of
Drinking
Water.

USEPA.
1987.
U.
S.
Environmental
Protection
Agency.
Carcinogenicity
assessment
of
aldrin
and
dieldrin.
Document
no.
EPA
600/
6­
87/
006,
August
1987.
Washington,
DC:
USEPA
Office
of
Health
and
Environmental
Assessment,
Carcinogenesis
Assessment
Group.

USEPA.
1980.
U.
S.
Environmental
Protection
Agency.
Ambient
water
quality
criteria
for
aldrin/
dieldrin.
Document
no.
EPA
440/
5­
80­
019.
Washington,
DC:
USEPA
Office
of
Water,
Office
of
Water
Regulations
and
Standards,
Criteria
and
Standards
Division.

Walker,
A.
I.
T.,
E.
Thorpe,
J.
Robinson,
and
M.
K.
Baldwin.
1971.
Toxicity
studies
on
the
photoisomerisation
product
of
dieldrin.
Meded.
Fac.
Landbouwwet.
Rijksuniv.
Gent.
36(
1):
398­
409
(
as
cited
in
IARC,
1974b;
IPCS,
1989).

Walker,
A.
I.
T.,
D.
E.
Stevenson,
J.
Robinson,
E.
Thorpe,
and
M.
Roberts.
1969.
The
toxicology
and
pharmacodynamics
of
dieldrin
(
HEOD):
Two­
year
oral
exposures
of
rats
and
dogs.
Toxicol.
Appl.
Pharmacol.
15:
345­
373
(
as
cited
in
ATSDR,
2000;
IPCS,
1989;
USEPA,
1980).

Walton,
M.
S.,
V.
Beck­
Bastone,
and
R.
L.
Baron.
1971.
Subchronic
toxicity
of
photodieldrin,
a
photodecomposition
product
of
dieldrin.
Toxicol.
Appl.
Pharmacol.
20(
1):
82­
88
(
as
cited
in
IPCS,
1989).
6­
29
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Dieldrin
 
February
2003
Winteringham,
F.
P.
W.
and
J.
B.
Barnes.
1955.
Comparative
response
of
insects
and
mammals
to
certain
halogenated
hydrocarbons
used
as
pesticides.
Physiol.
Rev.
35:
701
(
as
cited
in
USEPA,
1980).

Witherup,
S.,
K.
L.
Stemmer,
J.
L.
Roberts,
et
al.
1961.
Prolonged
cutaneous
contact
of
wool
impregnated
with
dieldrin.
The
Kettering
Laboratory
in
the
Department
of
Preventive
Medicine
and
Industrial
Health,
College
of
Medicine,
University
of
Cincinnati.
Cincinnati,
OH
(
as
cited
in
ATSDR,
2000).

Wolff,
T.,
E.
Deml,
and
H.
Wanders.
1979.
Aldrin
epoxidation,
a
highly
sensitive
indicator
specific
for
cytochrome
P­
450­
dependent
monooxygenase
activities.
Drug
Metab.
Dispos.
7:
301­
305
(
as
cited
in
ATSDR,
2000).

Wong,
D.
T.
and
L.
C.
Terriere.
1965.
Epoxidation
of
aldrin,
isodrin,
and
heptachlor
by
rat
liver
microsomes.
Biochem.
Pharmacol.
14:
375­
377
(
as
cited
in
ATSDR,
2000;
IARC,
1974a;
USEPA,
1980).

Wright,
A.
S.,
C.
Donninger,
R.
D.
Greenland,
K.
L.
Stemmer,
and
M.
R.
Zavon.
1978.
The
effects
of
prolonged
ingestion
of
dieldrin
on
the
livers
of
male
rhesus
monkeys.
Ecotoxicol.
Environ.
Saf.
1(
4):
477­
502
(
as
cited
in
IPCS,
1989;
TOXLINE,
2000c).
7­
1
Aldrin/
Dieldrin
 
February
2003
7.0
HAZARD
IDENTIFICATION
The
purpose
of
this
section
is
to
characterize
the
carcinogenic
and
non­
carcinogenic
health
effects
of
aldrin
and
dieldrin,
based
on
an
evaluation
of
information
from
both
human
epidemiological
and
case
studies
and
from
animal
studies.
In
addition,
mechanistic
studies
on
these
compounds
from
human,
animal,
and
in
vitro
experiments
are
reviewed,
and
possible
modes
of
action
for
some
of
their
various
non­
carcinogenic
and
carcinogenic
effects
are
discussed.

7.1
Human
Effects
This
section
briefly
highlights
the
rather
limited
number
of
human
case
and
epidemiological
studies
that
have
reported
acute
to
chronic
effects
resulting
from
exposure
to
aldrin
and/
or
dieldrin.

7.1.1
Short­
Term
Studies
The
short­
term
studies
summarized
below
primarily
reflect
the
oral
exposure
effects
of
aldrin
and
dieldrin
reported
in
humans
under
accidental
poisoning
scenarios.

General
Population
Aldrin
Jager
(
1970)
reported
the
acute
oral
lethal
dose
of
aldrin
in
an
adult
male
to
be
5.0
g
(
approximately
70
mg/
kg,
assuming
a
body
weight
of
70
kg).
A
somewhat
lower
ingested
dose
of
aldrin
(
25.6
mg/
kg)
has
been
reported
to
have
caused
convulsions
in
a
23­
year
old
male
after
20
minutes
(
Spiotta,
1951).
Although
his
convulsions
ceased
after
treatment
with
pentobarbital,
he
continued
to
exhibit
restlessness,
hypothermia,
tachycardia,
and
hypertension
for
up
to
5
days,
and
electroencephalogram
(
EEG)
abnormalities
for
up
to
6
months.

Severe
acute
intoxication
following
aldrin
exposure
in
humans
is
characterized
by
a
brief
period
of
excitation
or
drowsiness,
followed
by
convulsions,
muscle
twitching,
and
coma.
Hypothermia
generally
accompanies
death.
The
majority
of
individuals
intoxicated
with
aldrin,
however,
usually
regain
consciousness
and
recover
(
Hayes,
1982;
Jager,
1970).

Dieldrin
Dieldrin
has
been
reported
to
cause
hypersensitivity
and
muscular
fasciculations
that
may
be
followed
by
convulsive
seizures
and
associated
changes
in
the
EEG
pattern.
Acute
symptoms
of
intoxication
include
hyperirritability,
convulsions
and/
or
coma,
sometimes
accompanied
by
nausea,
vomiting
and
headache;
chronic
intoxication
may
result
in
fainting,
muscle
spasms,
tremors,
and
loss
of
weight.
The
lethal
dose
for
humans
is
estimated
to
be
about
5.0
g
(
ACGIH,
1984).
7­
2
Aldrin/
Dieldrin
 
February
2003
Black
(
1974)
observed
tachycardia,
elevated
blood
pressure,
and
convulsions
in
a
man
who
ingested
120
mg/
kg
dieldrin.
These
cardiovascular
effects
were
presumed
to
be
due
to
altered
activity
in
the
central
nervous
system
(
i.
e.,
increased
sympathetic
output),
as
the
symptoms
were
controlled
by
the
administration
of
$­
adrenergic
blocking
drugs.
Persistent
headaches,
irritability,
and
short­
term
memory
loss
were
also
reported
following
the
patient's
recovery
from
convulsions.

Sensitive
Populations
Children
are
generally
considered
at
greater
risk
than
adults
to
the
toxic
effects
of
chemicals
for
reasons
that
include
underdeveloped/
developing
organ
systems
or
capacities
(
e.
g.,
nervous
system,
digestive
and
reproductive
systems,
immune
systems,
metabolic
detoxication
capacity),
increased
potential
for
exposure,
increased
chemical
absorption,
etc.
One
study
reported
that
the
ingestion
of
approximately
120
mg
(
8.2
mg/
kg)
of
aldrin
by
a
3­
year
old
female
resulted
in
collapse
and
convulsions
within
5
minutes
and
death
within
12
hours
(
Hayes,
1982).

Garrettson
and
Curley
(
1969)
reported
convulsions
in
two
children
(
ages
2
and
4
years)
who
consumed
an
unknown
amount
of
a
5%
dieldrin
solution
(
also
containing
solvents
and
emulsifiers).
The
children
began
to
salivate
heavily,
and
then
developed
convulsions
within
15
minutes;
the
younger
child
died,
whereas
the
older
brother
had
liver
dysfunction
prior
to
recovering
completely.

7.1.2
Long­
Term
and
Epidemiological
Studies
The
long­
term
epidemiological
studies
were
conducted
mainly
in
populations
working
in
pesticide
manufacturing
plants,
although
some
utilized
volunteers.
In
most
cases,
some
combination
of
oral,
inhalation,
and
dermal
routes
of
exposure
were
probably
involved.

General
Populations
Aldrin
One
male,
employed
21
years
at
a
chemical
plant
and
reassigned
to
the
handling
of
aldrin
concentrate
(
period
and
levels
of
exposure
were
not
specified),
experienced
involuntary
jerking
(
rapid
flexor
movement)
of
his
hands
and
forearms,
vomiting,
and
chronic
irritability,
and
insomnia
(
Hodge
et
al.,
1967).
His
EEG
showed
alpha­
wave
irregularities,
with
discharges
of
slow
and
sharp
waves.
After
exposure
to
aldrin
was
discontinued,
his
condition
rapidly
improved.

Dieldrin
(
mean
=
13
ng/
g
whole
milk)
was
found
in
the
breast
milk
of
women
whose
homes
were
treated
annually
(
or
more
frequently)
with
organochlorine
pesticides
(
Stacey
and
Tatum,
1985).
A
correlation
between
dieldrin
levels
in
the
milk
and
aldrin
treatment
of
homes
was
observed.
Dieldrin
levels
in
breast
milk
rose
until
the
seventh
or
eighth
month
after
treatment
of
homes
was
discontinued.
No
data
were
provided
on
the
health
effects
of
children
exposed
to
dieldrin­
contaminated
breast
milk.
7­
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Aldrin/
Dieldrin
 
February
2003
Edwards
and
Priestly
(
1994)
reported
elevated
plasma
dieldrin
levels
and
hepatic
enzyme
activity
(
as
measured
by
urinary
D­
glucaric
acid
excretion)
in
33
workers
(
29
males
and
4
females)
from
2
south
Australian
suburban
pesticide
treatment
businesses;
they
had
worked
in
the
industry
ranging
from
3
months
to
20
years.
The
plasma
dieldrin
concentrations
in
workers
applying
aldrin
ranged
from
2.5
to
250
ng/
mL,
while
those
for
workers
not
involved
with
aldrin
exposure
(
office
staff)
had
dieldrin
levels
ranging
from
0.7
to
26
ng/
mL.
However,
there
was
no
correlation
between
high
D­
glucaric
acid
excretion
and
plasma
dieldrin
levels.

Dieldrin
Hunter
and
Robinson
(
1967)
observed
no
effect
on
central
nervous
system
activity
(
as
measured
by
EEG),
peripheral
nerve
activity,
or
muscle
activity
in
volunteers
administered
dieldrin
daily
for
18
months
at
doses
as
high
as
0.003
mg/
kg
bw/
day.

Aldrin/
Dieldrin
No
increase
in
mortality
from
any
cause
was
reported
in
workers
(
n
=
233)
who
had
been
employed
in
the
manufacture
of
aldrin,
dieldrin,
and
other
pesticides
at
a
facility
in
the
Netherlands
for
more
than
4
years
(
Van
Raalte,
1977;
Versteeg
and
Jager,
1973).

Subsequent
studies
conducted
from
the
Netherlands
included
several
years
of
follow
up,
which
may
be
summarized
as
follows.
De
Jong
(
1991)
reported
mortality
data
in
a
20­
year
follow­
up
study
of
cohorts
exposed
to
insecticides
for
at
least
1
year
between
1954
and
1970
(
total
cohorts
=
570
workers).
At
the
time
of
the
vital
status
cut­
off
date
(
January
1,
1987),
of
these
570
workers,
445
(
78%)
were
alive;
76
(
13.3%)
were
deceased;
34
(
6.0%)
emigrated;
and
15
(
2.6%)
were
lost
to
the
follow
up.
Workers
on
the
study
represented
14,740
person
years
of
observation.
Exposure
estimates
were
made
based
on
the
available
blood
dieldrin
data
collected
from
343
of
the
workers.
The
workers
were
divided
into
low,
medium,
and
high
exposure
categories
having
estimated
mean
daily
aldrin/
dieldrin
intakes
of
90,
419,
and
1019
:
g,
respectively
(
corresponding
mean
lifetime
intake
values
were
88,
419,
and
1704
mg).
The
standardized
mortality
ratios
(
SMRs)
for
all
causes
of
death
for
the
workers
exposed
to
aldrin/
dieldrin,
as
compared
to
Netherlands
national
mortality
rates,
were
80.6,
86.8,
and
68.9
for
low,
moderate,
and
high
exposures,
respectively.

A
more
recent
study
from
the
Netherlands
reported
on
the
mortality
of
the
same
cohorts
with
a
latter
follow­
up
date
(
de
Jong
et
al.,
1997).
Of
the
570
workers,
70.5%
(
402)
were
alive;
20.7%
were
deceased
(
118);
6.2%
(
35)
emigrated,
and
2.6%
(
15)
were
lost
to
the
follow­
up,
at
the
cut
off
date
of
January
1,
1993.
The
total
mortality
observed
from
all
causes
of
death
in
all
the
cohorts
was
lower
than
the
expected
number
of
deaths,
calculated
from
national
data
according
to
age,
period,
and
causes
of
specific
mortalities
(
118
deaths
observed
versus
156
deaths
expected;
SMR
=
75.6,
with
a
95%
confidence
interval
of
63
to
91).
Similar
lower
trends
were
observed
for
mortality
rates
from
cardiovascular
disease
and
non­
malignant
respiratory
disease.
Of
all
the
types
of
cancers,
only
two
(
rectum
and
liver)
had
higher
frequency
than
expected,
but
these
results
were
not
dose
dependent.
Six
deaths
from
rectal
cancer
were
observed
in
the
cohorts,
as
compared
to
1.5
expected
(
SMR
=
390.4,
with
a
95%
confidence
7­
4
Aldrin/
Dieldrin
 
February
2003
interval
of
143
to
850).
Two
deaths
due
to
liver
cancer
were
observed
in
the
cohorts,
as
compared
to
0.9
expected
(
SMR
=
225,
with
a
95%
confidence
interval
of
27
to
813).
Stratification
of
the
data
according
to
the
type
of
job
(
operators,
maintenance
workers,
supervisors)
showed
a
significantly
(
p
<
0.01)
increased
mortality
rate
for
rectal
cancer
only
in
the
operator
group
(
de
Jong
et
al.,
1997).

Three
follow­
up
cohort
studies
were
reported
on
the
mortality
rates
of
workers
from
a
pesticide
manufacturing
plant
in
Denver,
CO
(
Ditraglia
et
al.,
1981;
Brown,
1992;
Amaoteng­
Adjepong
et
al.,
1995).
In
the
first
retrospective
cohort
study,
Ditraglia
et
al.
(
1981)
reported
SMRs
for
1,155
workers
who
had
been
employed
at
the
plant
for
at
least
6
months
prior
to
1964
and
were
exposed
to
aldrin/
dieldrin.
Of
the
1,155
workers,
75%
(
870)
were
alive,
15%
(
173)
deceased,
and
10%
(
112)
were
of
undetermined
vital
status
as
of
the
study
cut­
off
date
(
December
31,
1976).
Workers
in
the
study
represented
24,939
person
years
of
observation.
They
were
mainly
white
males,
and
the
mortality
rates
of
the
exposed
population
were
compared
to
white
male
cause­
specific
mortality
rates
in
the
U.
S.
The
mortality
rate
for
all
causes
of
death
(
combination
of
malignant,
circulatory,
nonmalignant
respiratory,
and
nervous
system
diseases)
was
significantly
lower
in
the
exposed
group
than
in
the
controls
(
SMR
=
84,
with
a
95%
confidence
interval
of
72
to
98).
The
SMRs
for
neoplasms
of
the
liver
and
the
lymphatic/
hematopoietic
system
were
not
statistically
different
from
100,
the
values
corresponding
to
225
(
95%
confidence
interval
of
39
to
1267)
and
147
(
95%
confidence
interval
of
54
to
319),
respectively.
However,
the
authors
reported
a
significant
increase
in
the
SMR
for
nonmalignant
respiratory
disease
at
212,
with
a
95%
confidence
interval
of
133
to
320.

The
study
by
Brown
(
1992)
extended
the
observations
reported
by
Ditraglia
et
al.
(
1981)
having
a
study
cut­
off
date
of
December
31,
1987,
and
1158
workers.
Of
these,
803
(
70%)
were
alive,
337
(
29%)
were
deceased,
and
13
(
10%)
were
of
undetermined
vital
status.
Workers
in
the
study
represented
34,479
person
years
of
observation.
The
mortality
rate
for
all
causes
of
death
(
combination
of
malignant,
circulatory,
nonmalignant
respiratory,
and
cerebrovascular
diseases)
was
lower
for
the
exposed
cohort
than
for
controls
(
SMR
=
87,
with
a
95%
confidence
interval
of
78
to
97).
Comparing
the
cohort
mortality
rate
to
national,
state,
or
county
statistics
did
not
affect
the
SMR
for
all
causes
of
death.
However,
the
SMR
for
liver/
biliary
cancer
was
higher
than
expected,
with
values
corresponding
to
393
(
CI
=
127
to
920),
510
(
CI
=
165
to
1,191),
or
486
(
CI
=
157
to
1,136)
when
compared
to
U.
S.,
state,
or
county
mortality
rates,
respectively.
Of
the
five
observed
cases
of
liver
and
biliary
cancer,
two
were
also
in
the
dibromochloropropane
(
DBCP)
registry.

Finally,
Amaoteng­
Adjepong
et
al.
(
1995)
updated
the
SMRs
of
the
workers
in
the
Denver
pesticide
manufacturing
plant;
the
study
population
is
similar
to
those
mentioned
in
the
previous
two
studies
(
Ditraglia
et
al.
1981;
Brown,
1992),
except
that
most
of
the
employees
worked
at
the
plant
between
1952
and
1982,
and
the
race
was
reported
for
most
(
n
=
2,384).
The
unknown
race
of
262
workers
was
classified
as
white.
The
cohort
had
the
vital
status
of
1,764
(
74%),
496
(
21%),
and
124
(
5.0%)
for
living,
deceased,
and
unknown
categories,
at
the
time
of
the
cut
off
date
of
January
1,
1991.
Within
the
cohort,
87%
of
the
workers
consisted
of
white
males
(
n
=
2,072);
10%
were
white
females
(
234);
3%
were
black
males
(
n
=
68);
and
<
1%
were
black
females
(
n
=
10).
The
analysis
of
the
data
suggested
no
positive
relationship
between
7­
5
Aldrin/
Dieldrin
 
February
2003
aldrin/
dieldrin
exposure
and
mortality
due
to
liver
cancer
or
other
causes
of
death
(
respiratory,
circulatory,
or
nervous
system
diseases).

Nair
et
al.
(
1992)
reported
finding
aldrin
and
dieldrin
levels
in
adipose
tissue,
breast
milk,
and
serum
samples
collected
from
Delhi
female
residents
(
18
to
24
years
old;
n
=
12)
during
1989
through
1990.
The
subjects
were
from
low
socioeconomic
status
and
were
residing
in
parts
of
Delhi
exposed
to
severe
automobile
pollution.
The
average
aldrin
concentrations
were
0.048,
0.003,
and
0.004
ppb
in
adipose
tissue,
breast
milk,
and
serum,
respectively,
and
the
corresponding
average
dieldrin
concentrations
were
0.099,
0.06,
and
0.002
ppb,
suggesting
greater
concentration
of
aldrin/
dieldrin
in
adipose
tissues.
A
significant
correlation
was
reported
between
the
levels
of
aldrin/
dieldrin
found
in
adipose
tissue
and
those
found
in
serum
(
p<
0.01;
r
=
0.503).
The
authors
also
observed
that
aldrin
and
dieldrin
values
were
higher
in
the
breast
milk
of
primagravidae
(
first
time
deliverers)
when
compared
to
women
who
had
undergone
their
second
delivery.
They
concluded
that
the
aldrin/
dieldrin
levels
in
Delhi
residents
were
low
when
compared
to
the
values
found
in
populations
from
developed
countries.
Conflicting
reports
exist
on
the
effect
of
aldrin/
dieldin
on
hematological
parameters.
A
farmer
with
multiple
exposures
to
insecticides
that
contained
dieldrin
died
in
a
hemolytic
crisis
after
developing
immunohemolytic
anemia
(
Muirhead
et
al.,
1959).
Immunologic
testing
revealed
a
strong
antigenic
response
to
red
blood
cells
coated
with
dieldrin.
In
another
study,
a
worker
from
an
orange
grove
developed
aplastic
anemia
and
died
following
repeated
exposures
to
aldrin
during
spraying
(
Pick
et
al.,
1965).
In
the
latter
study,
the
relationship
between
aldrin
exposure
and
the
aplastic
anemia
was
considerably
more
tenuous,
being
linked
only
in
that
the
onset
of
symptoms
corresponded
with
spraying
and
the
condition
deteriorated
upon
subsequent
exposure.
However,
in
another
study
of
workers
employed
in
a
pesticide
manufacturing
plant
for
4
years,
no
abnormal
values
for
hemoglobin,
white
blood
cells,
or
erythrocyte
sedimentation
rate
were
found
(
Jager,
1970).
Further,
workers,
who
had
been
involved
in
either
the
manufacture
or
application
of
pesticides
and
who
had
elevated
blood
dieldrin
levels,
had
no
hematological
effects
of
clinical
significance
(
Warnick
and
Carter,
1972).

Sensitive
Populations
No
long­
term
studies
were
located
that
examined
the
adverse
health
effects
of
aldrin
or
dieldrin
exposure
in
children
(
who
in
general
are
considered
to
be
among
the
most
sensitive
populations
for
exposure
to
chemicals),
or
in
any
other
potentially
high­
risk
population
(
e.
g.,
the
aged
or
those
with
pre­
existing
liver
or
neurological
disease).

7.2
Animal
Studies
7.2.1
Acute
Toxicity
(
Oral,
Inhalation,
Dermal)

Oral
Exposure
The
oral
median
lethal
dose
(
LD
50)
values
for
aldrin
in
laboratory
animals
are
as
follows:
mice,
44
mg/
kg
bw
(
purity
not
reported;
Borgmann
et
al.,
1952);
rats,
39
to
60
mg/
kg
bw
(
purity
not
reported;
Gaines,
1969);
guinea
pigs,
33
mg/
kg
bw
(
purity
not
reported;
Borgmann
et
al.,
7­
6
Aldrin/
Dieldrin
 
February
2003
1952);
female
rabbits,
50
to
80
mg/
kg
bw
(
purity,
95%;
Treon
and
Cleveland,
1955);
and
dogs,
65
to
95
mg/
kg
bw
(
purity
not
reported;
Borgmann
et
al.,
1952).

The
doses
at
which
aldrin
is
acutely
lethal
in
experimental
animals
are
quite
similar
to
those
for
dieldrin.
Oral
LD
50
values
for
single
doses
of
aldrin
in
rats
ranged
from
39
to
64
mg/
kg
bw
(
Gaines,
1960;
Treon
et
al.,
1952),
while
those
for
single
doses
of
dieldrin
ranged
from
37
to
46
mg/
kg
bw
(
Gaines,
1960;
Lu
et
al.,
1965;
Treon
et
al.,
1952).
Aldrin
was
lethal
in
female
rats
at
a
slightly
lower
dose
when
it
was
administered
in
solution
in
oil
(
LD
50
=
48
mg/
kg
bw),
than
when
it
was
administered
in
a
kerosene
vehicle
(
LD
50
=
64
mg/
kg
bw)
(
Treon
et
al.,
1952).

The
age
of
the
animals
appeared
to
influence
the
acute
toxicity
of
a
single
administration
of
dieldrin.
Two
week­
old
rats
had
an
LD
50
of
25
mg/
kg
bw,
which
is
lower,
as
expected,
than
the
LD
50
value
(
37
mg/
kg
bw)
found
in
young
adult
rats
(
Lu
et
al.,
1965).
However,
newborn
rats
had
a
relatively
high
LD
50
of
168
mg/
kg
bw
(
Lu
et
al.,
1965).

Acute
toxicity
in
animals
is
characterized
by
increased
irritability,
salivation,
hyperexcitability,
tremors
followed
by
clonic/
tonic
convulsions,
anorexia
and
loss
of
body
weight,
depression,
prostration,
and
eventual
death
(
Borgmann
et
al.,
1952;
Hodge
et
al.,
1967).

Inhalation
Exposure
Treon
et
al.
(
1957)
exposed
cats,
guinea
pigs,
rats,
rabbits,
and
mice
to
aldrin
vapors
and
particles
generated
by
sublimating
aldrin
at
200
°
C.
Aldrin
levels
of
108
mg/
m3
for
1
hour
resulted
in
the
death
of
9
out
of
10
rats,
3
out
of
4
rabbits,
and
2
out
of
10
mice.
Cats
and
guinea
pigs
were
less
sensitive.
One
out
of
one
cat
and
no
guinea
pigs
died
following
exposure
to
215
mg/
m3
for
4
hours.
Interpretation
of
the
results
of
this
study
is
limited
in
that
sublimation
may
have
resulted
in
the
generation
of
atmospheres
containing
a
higher
proportion
of
volatile
contaminants
than
would
be
expected
in
atmospheres
typical
of
most
occupational
exposures.

Dermal
Exposure
In
rats,
a
single
dermal
application
of
aldrin
in
xylene
produced
an
LD
50
value
of
60
mg/
kg
bw
in
female
rats
and
90
mg/
kg
bw
in
male
rats
(
Gaines,
1960).
A
single
24­
hour
dermal
exposure
of
rabbits
to
dry
crystallized
aldrin
or
dieldrin
resulted
in
LD
50
values
between
600
and
1,250
mg/
kg
bw
for
both
chemicals
(
Treon
et
al.,
1953).

7.2.2
Short­
Term
Studies
Oral
Exposure
In
a
short­
term
study,
Treon
and
Cleveland
(
1955)
observed
100%
mortality
within
2
weeks
in
groups
of
male
and
female
Carworth
rats
(
total
number
and
number/
sex
not
reported)
that
were
fed
aldrin
(
purity
95%)
at
a
concentration
of
300
ppm
(
an
approximate
dose
of
15
mg/
kg
bw/
day,
based
on
Lehman,
1959).
No
mortality
was
noted
at
lower
doses.
Administration
of
a
diet
containing
25
ppm
aldrin
(
purity
95%),
an
approximate
dose
of
0.625
mg/
kg
bw/
day,
to
7­
7
Aldrin/
Dieldrin
 
February
2003
2
male
and
3
female
beagle
dogs
induced
fatalities
after
periods
of
feeding
ranging
from
9
to
15
days
(
Treon
and
Cleveland,
1955).

Kolaja
et
al.
(
1996a)
investigated
the
short­
term
effects
in
male
Fisher
344
rats
and
B6C3F
1
mice
(
5
animals/
species/
group)
after
administration
of
dieldrin
at
0
(
control),
0.1,
1.0,
3.0,
or
10.0
mg/
kg
bw
diet
for
7
or
14
days
(
approximate
doses
in
rats
of
0.005,
0.05,
0.15,
or
0.5
mg/
kg
bw/
day,
and
in
mice
of
0.015,
0.15,
0.45,
or
1.5
mg/
kg
bw/
day;
based
on
Lehman,
1959).
Relative
liver
weights
(
liver
weight/
body
weight)
in
mice
were
significantly
increased
at
all
doses
tested
compared
to
controls.
However,
in
rats,
apparent
increases
in
relative
liver
weights
were
found
only
in
the
10.0
mg/
kg
bw
diet
dieldrin
group
after
7
days
of
treatment.
Dieldrin
was
not
severely
hepatotoxic
in
either
species,
as
evidenced
by
no
changes
in
the
activities
of
serum
enzymes
such
as
ALT
and
AST,
and
no
histopathology.

In
an
another
study,
Kolaja
et
al.
(
1996b)
reported
selective
promotion
of
hepatic
focal
lesions
in
male
B6C3F
1
mice,
but
not
in
male
Fisher
344
rats,
following
administration
of
dieldrin
at
0.1,
1.0,
or
10.0
mg/
kg
bw
diet
(
5
animals/
group)
for
7
days
(
approximate
doses
in
rats
of
0.005,
0.05,
or
0.5,
mg/
kg
bw/
day,
and
in
mice
of
0.015,
0.15,
or
1.5
mg/
kg
bw/
day;
based
on
Lehman,
1959).
Study
animals
including
controls
were
injected
intraperitoneally
with
the
hepatic
carcinogen,
diethyl
nitrosamine
(
150
mg/
kg
bw/
week,
2x
for
rats;
25
mg/
kg
bw/
week,
8x
for
mice),
prior
to
dieldrin
treatment
in
order
to
enhance
the
formation
of
hepatic
lesions.
No
significant
effects
on
the
number
or
volume
of
hepatic
focal
lesions
(
total),
DNA
labeling
index,
or
relative
liver
weight
(
liver
to
body
weight
ratio)
were
observed
in
the
rats.
However,
significant
increases
(
p
<
0.05)
in
the
number
of
hepatic
focal
lesions
and
in
hepatic
focal
lesion
volume,
DNA
labeling
index,
and
relative
liver
weight
were
noted
in
the
mice
treated
with
the
high
dose
of
dieldrin.
No
changes
in
body
weight
or
in
the
apoptotic
index
of
the
hepatic
focal
lesions
were
observed
at
any
of
the
doses
tested,
either
in
rats
or
mice.

Inhalation
Exposure
No
studies
were
obtained
that
investigated
the
short­
term
toxic
effects
of
aldrin
or
dieldrin
in
animals
after
inhalation
exposure.

Dermal
Exposure
No
studies
were
obtained
that
investigated
the
short­
term
toxic
effects
of
aldrin
or
dieldrin
in
animals
after
dermal
exposure.

7.2.3
Subchronic
Studies
Oral
Exposure
Aldrin
Decreased
body
weight
gain
and
increased
mortality
were
observed
in
the
high­
dose
group
of
Osborne­
Mendel
rats
(
5/
sex/
group)
fed
aldrin
(
technical
grade,
95%
pure)
in
the
diet
at
7­
8
Aldrin/
Dieldrin
 
February
2003
concentrations
of
0,
40,
80,
160,
or
320
ppm
aldrin
(
doses
of
0
and
approximately
2,
4,
8,
or
16
mg/
kg
bw/
day,
respectively,
based
on
a
food
consumption
factor
of
0.05
from
Lehman,
1959)
for
42
days
then
and
observed
for
an
additional
14
days
(
NCI,
1978).
The
No­
Observed­
Adverse­
Effect
Level
(
NOAEL)
for
this
study
was
160
ppm
(
8
mg/
kg
bw/
day).
This
study
was
a
range­
finding
study
for
a
long­
term
carcinogenicity
study;
therefore,
a
complete
toxicology
profile
was
not
obtained
(
e.
g.,
biochemical
and
hematology
assessments
were
not
performed).

In
groups
of
B6C3F
1
mice
(
5/
sex/
group)
fed
aldrin
(
technical
grade,
95%
pure)
at
concentrations
of
0,
2.5,
5,
10,
20,
40,
or
80
ppm
(
doses
of
0
and
approximately
0.375,
0.75,
1.5,
3,
6,
or
12
mg/
kg
bw/
day,
respectively,
based
on
a
food
consumption
factor
of
0.15
from
Lehman,
1959)
in
the
diet
for
42
days,
100%
mortality
was
observed
in
the
40
and
80
ppm
(
6
and
12
mg/
kg
bw/
day,
respectively)
groups.
One
male
and
one
female
died
in
the
20
ppm
(
3
mg/
kg
bw/
day)
group;
10
and
20
ppm
(
1.5
and
3
mg/
kg
bw/
day,
respectively)
were
therefore
considered
the
NOAEL
and
Lowest­
Observed­
Adverse­
Effect
Level
(
LOAEL)
values,
respectively,
for
this
study
(
NCI,
1978).
This
study
was
a
range­
finding
study
for
a
long­
term
carcinogenicity
study;
therefore,
a
complete
toxicology
profile
was
not
obtained.

Dieldrin
Kolaja
et
al.
(
1996a)
reported
no
statistically
significant
differences
in
either
body
weight
gains,
food
consumption,
or
water
consumption
in
male
B6C3F
1
mice
or
Fisher
344
rats
that
were
administered
dieldrin
at
concentrations
of
0.1,
1.0,
3.0,
or
10.0
mg/
kg
bw
diet
for
21,
28,
or
90
days
(
approximate
doses
in
rats
of
0.005,
0.05,
0.15,
or
0.5
mg/
kg
bw/
day,
respectively,
and
in
mice
of
0.015,
0.15,
0.45,
or
1.5
mg/
kg
bw/
day,
respectively;
based
on
Lehman,
1959).
Also,
no
severe
hepatotoxicity
was
observed
in
dieldrin­
treated
animals,
as
evidenced
by
no
changes
in
activities
of
the
serum
enzymes
ALT
(
alanine
aminotransferase)
and
AST
(
aspartamine
aminotransferase),
and
no
apparent
histopathology.
However,
relative
liver
weights
(
liver/
body
weight
ratios)
were
significantly
increased
in
mice
(
but
not
in
rats)
at
the
highest
dose
tested.

In
a
subsequent
report,
Kolaja
et
al.
(
1996b)
reported
that
dieldrin
administered
to
groups
of
male
B6C3F
1
mice
or
Fisher
344
rats
(
5/
group/
species/
dose)
at
concentrations
of
0.1,
1.0,
or
10.0
mg/
kg
bw
diet
for
30
or
60
days
(
approximate
doses
in
rats
of
0.005,
0.05,
or
0.5
mg/
kg
bw/
day,
respectively,
and
in
mice
of
0.015,
0.15
or
1.5
mg/
kg
bw/
day,
respectively;
based
on
Lehman,
1959)
caused
the
selective
promotion
of
hepatic
focal
lesions
in
the
mice
but
not
in
the
rats.
Study
animals,
including
controls,
were
injected
intraperitoneally
with
the
hepatic
carcinogen,
diethyl
nitrosamine
(
150
mg/
kg
bw/
week,
2x
for
rats;
25
mg/
kg
bw/
week,
8x
for
mice),
prior
to
dieldrin
treatment
in
order
to
enhance
the
formation
of
hepatic
lesions.
No
significant
effects
on
the
number
or
volume
of
hepatic
focal
lesions
(
total)
were
observed
for
rats
at
any
of
the
doses
tested
during
the
30
or
60
days
after
dieldrin
treatment.
However,
significant
increases
(
p
<
0.05)
in
the
number
of
hepatic
focal
lesions
and
in
hepatic
focal
lesion
volume
and
DNA
labeling
index
were
noted
in
mice
treated
with
the
high
dose
of
dieldrin
after
30
and
90
days.
Dieldrin
treatment
also
caused
an
inconsistent
increase
in
relative
liver
weights
in
both
rats
and
mice.
Changes
in
body
weight
or
in
the
apoptotic
index
of
hepatic
focal
lesions
were
not
observed
at
any
dose
or
duration
tested,
in
either
rats
or
mice.
7­
9
Aldrin/
Dieldrin
 
February
2003
Stevenson
et
al.
(
1995)
also
reported
that
dieldrin
caused
an
increase
in
hepatotoxicity
such
as
liver
enlargement,
increased
DNA
synthesis
in
hepatocytes,
hypertrophy
of
centrilobular
hepatocytes,
and
induction
of
hepatic
ethoxyreosrufin
0­
deethylase
(
microsomal
mixed
function
oxidase)
activity
at
the
highest
dose
in
male
B6C3F
1
mice
fed
with
dieldrin
at
1,
3,
or
10
mg/
kg
diet
for
28
days
(
approximate
doses
of
0.15,
0.45,
or
1.5
mg/
kg
bw/
day;
based
on
Lehman,
1959).

Inhalation
Exposure
No
studies
were
obtained
that
investigated
the
toxic
effects
of
aldrin
or
dieldrin
in
animals
after
subchronic
inhalation
exposure.

Dermal
Exposure
Aldrin
or
dieldrin
(
dry
powder)
applied
to
rabbit
skin
for
2
hours/
day,
5
days/
week
for
10
weeks,
was
reported
to
have
had
no
discernible
effects
(
IPCS,
1989).

7.2.4
Neurotoxicity
Oral
Exposure
Aldrin
Paul
et
al.
(
1992)
reported
behavioral
impairments
in
Wistar
rats
(
10/
group)
that
were
administered
1
mg/
kg
bw/
day
aldrin
(
technical
grade,
90%
pure)
by
gavage
for
up
to
90
days.
Aldrin
inhibited
muscle
coordination
(
measured
using
rota­
rod
apparatus)
beginning
on
the
15th
day
in
both
sexes,
with
greater
motor
deterioration
occurring
in
males.
Aldrin
also
inhibited
learning
ability
and
the
conditioned
avoidance
response
(
measured
in
a
pole­
climbing
apparatus),
as
the
number
of
animals
responding
to
simultaneous
unconditioned
and
conditioned
stimuli
was
significantly
reduced
(
p
<
0.05)
in
aldrin­
treated
groups
when
compared
to
controls.

Neurotoxic
signs
observed
in
cattle
poisoned
with
unspecified
dietary
concentrations
of
aldrin
included
tremors,
running,
hyperirritability,
and
seizures
(
Buck
and
Van
Note,
1968).
Casteel
et
al.
(
1993)
reported
neurological
and
muscular
symptoms,
such
as
ataxia,
tremors,
hypersalivation,
diarrhea,
and
disorientation
in
6
calves;
lateral
recumbency
and
intermittent
tonoclonic
convulsions
in
2
calves;
and
severe
signs
such
as
death
in
10
calves,
in
a
group
of
feedlot
cattle
(
n
=
90)
exposed
to
aldrin­
contaminated
feed
in
northwest
Missouri.
The
selffeeders
in
the
feedlot
contained
from
54
to
528
:
g
aldrin/
g
of
feed.
Analysis
of
aldrin
and
dieldrin
in
the
rumen
content
of
two
dead
calves
revealed
the
concentrations
of
aldrin
as
20.6
and
22.4
:
g/
g
of
ingesta.
The
mean
dieldrin
concentrations
in
fat
samples
that
were
collected
50
days
after
the
withdrawal
of
contaminated
feed
from
the
calves
ranged
from
9.7
to
18.8
:
g/
g,
and
the
approximate
half­
lives
of
dieldrin
in
the
adipose
tissue
of
calves
ranged
from
53
to
231
days.
7­
10
Aldrin/
Dieldrin
 
February
2003
Dieldrin
Convulsions
were
observed
in
rats
given
single
doses
of
dieldrin
ranging
from
40
to
50
mg/
kg
(
Wagner
and
Greene,
1978;
Woolley
et
al.,
1985).
Transient
hypothermia
and
anorexia
were
also
reported
following
a
single
dose
of
40
mg/
kg
(
Woolley
et
al.,
1985).
Tremors
were
observed
in
rats
receiving
a
dose
of
0.5
mg/
kg
bw/
day
for
60
days
(
Mehrotra
et
al.,
1988),
and
hyperexcitability
was
observed
with
dieldrin
at
2.5
mg/
kg
bw/
day
in
an
8­
week
study
(
Wagner
and
Greene
1978).
Cerebral
edema
and
small
foci
of
degeneration
were
reported
in
rats
exposed
to
dieldrin
at
0.016
mg/
kg
bw/
day
for
2
years
(
Harr
et
al.,
1970),
although
the
study
had
various
limitations.

Operant
behavior
was
reported
to
have
been
disrupted
in
rats
following
single
doses
of
dieldrin
ranging
from
0.5
to
16.7
mg/
kg
(
Burt,
1975;
Carlson
and
Rosellini,
1987).
A
lower
dose
of
dieldrin
(
0.025
mg/
kg
bw/
day)
for
a
longer
duration
(
60
to
120
days)
was
also
observed
to
impair
operant
behavior
in
rats
(
Burt,
1975).

EEGs
taken
from
dogs
exposed
to
dieldrin
at
0.05
mg/
kg
bw/
day
for
2
years
were
normal
(
Walker
et
al.
1969).
However,
dogs
were
reported
to
develop
convulsions
when
given
0.5
mg/
kg
bw/
day
for
25
months
(
Fitzhugh
et
al.
1964).

Aldrin/
Dieldrin
When
aldrin
or
dieldrin
was
administered
to
rats
for
3
days,
convulsions
were
observed
at
a
dose
of
10
mg/
kg
bw/
day
(
Mehrotra
et
al.,
1989).
Histopathological
changes
were
found
in
the
brain
cells
of
rats
that
were
exposed
for
6
months
to
2.75
mg/
kg
bw/
day
of
either
aldrin
or
dieldrin
(
Treon
et
al.,
1951a).

Irritability,
tremors,
and
convulsions
were
observed
in
rats
fed
aldrin/
dieldrin
at
dietary
concentrations
ranging
from
0.1
to
65
ppm
in
several
1.5­
to
2.5­
year
studies
(
Deichmann
et
al.,
1970;
NCI,
1978;
Walker
et
al.,
1969).
Hyperexcitability
was
observed
in
Osborne­
Mendel
rats
exposed
for
74
to
80
weeks
to
aldrin
in
the
diet
at
30
and
60
ppm
(
approximate
doses
of
1.5
and
3.0
mg/
kg
bw/
day,
respectively,
according
to
Lehman,
1959)
(
NCI,
1978),
as
were
tremors
and
clonic
convulsions
after
31
months
exposure
to
20,
30,
or
50
ppm
(
approximate
doses
of
1.0.,
1.5,
or
2.5
mg/
kg
bw/
day,
respectively)
(
Deichmann
et
al.,
1970).
Similarly,
hyperexcitability
was
observed
in
Osborne­
Mendel
rats
fed
29
ppm
dieldrin
for
80
weeks
or
65
ppm
for
59
weeks
(
approximate
doses
of
1.45
and
3.25
mg/
kg
bw/
day,
respectively)
(
NCI,
1978).
In
a
companion
study
(
NCI,
1978),
Fischer
344
rats
that
were
fed
dieldrin
for
2
years
at
2,
10,
or
50
ppm
(
approximate
doses
of
0.1,
0.5,
or
2.5
mg/
kg
bw/
day,
respectively)
showed
convulsions,
tremors,
and
nervous
behavior
at
the
high
dose.
CF
rats
fed
0.1,
1,
or
10
ppm
dieldrin
(
approximate
doses
of
0.005,
0.05,
or
0.5
mg/
kg
bw/
day,
respectively)
for
2
years
displayed
irritability,
tremors,
and
convulsions
(
Walker
et
al.,
1969);
the
latter
2
effects
were
also
noted
in
Osborne­
Mendel
rats
exposed
to
20,
30,
or
50
ppm
dieldrin
(
approximate
doses
of
1,
1.5,
or
2.5
mg/
kg
bw/
day,
respectively)
for
29
months
(
Deichmann
et
al.,
1970).
7­
11
Aldrin/
Dieldrin
 
February
2003
B6C3F
1
mice
showed
slightly
greater
sensitivity
than
did
the
rats
in
the
NCI
(
1978)
80­
week
bioassays,
with
hyperexcitability
observed
at
dietary
exposures
of
aldrin
as
low
as
3
ppm
(
females)
and
4
ppm
(
males)
(
approximate
doses
of
0.45
and
0.60
mg/
kg
bw/
day,
respectively,
according
to
Lehman,
1959);
and
with
hyperexcitability,
tremors,
and
hyperactivity
observed
at
dietary
exposures
of
dieldrin
as
low
as
2.5
ppm
for
both
sexes
(
approximate
dose
of
0.38
mg/
kg
bw/
day).

Dogs
given
aldrin
at
0.89
to
1.78
mg/
kg
bw/
day,
or
dieldrin
at
0.73
to
1.85
mg/
kg
bw/
day,
for
up
to
9
months
experienced
neuronal
degeneration
in
the
cerebral
cortex
and
convulsions
(
Treon
et
al.,
1951b).
At
these
doses,
aldrin­
treated
dogs
also
displayed
hypersensitivity
to
stimulation,
twitching,
and
tremors,
while
at
higher
doses,
the
basal
ganglia
and
cerebellum
were
reported
to
exhibit
degenerative
changes.

Inhalation
Exposure
No
studies
were
obtained
that
investigated
the
neurotoxic
effects
in
animals
resulting
from
inhalation
exposure
to
either
aldrin
or
dieldrin.

Dermal
Exposure
In
a
study
examining
the
effects
of
acute
dermal
exposure
to
aldrin
or
dieldrin,
Treon
et
al.
(
1953)
reported
the
induction
of
tremors
and
convulsions
in
rabbits.
However,
the
doses
associated
with
these
effects
were
not
reported.

Other
Routes
of
Exposure
Castro
et
al.
(
1992)
reported
the
effects
of
prenatal
exposure
to
aldrin
on
the
behavioral
development
of
90
day­
old
adult
rats.
Pregnant
female
rats
(
10
to
20/
group)
were
subcutaneously
injected
with
either
aldrin
(
1.0
mg/
kg
bw)
or
its
vehicle
(
0.9%
sodium
chloride
plus
Tween
80)
from
day
1
of
pregnancy
until
delivery.
Prenatal
exposure
to
aldrin
reportedly
produced
no
changes
at
90
days
in
aldrin
or
dieldrin
levels
in
serum,
or
in
cellular
and
structural
organization
of
cerebral
cortex
neurons,
or
in
the
adult
animals'
behavior
as
determined
by
an
avoidance
learning
test.
However,
prenatal
administration
of
aldrin
was
found
to
produce
a
significant
increase
(
p
<
0.05)
in
the
locomotor
frequency
of
experimental
rats
at
21
and
90
days.
Also,
the
performance
of
adult
rats
in
the
hole­
board
apparatus
(
total
number
and
duration
of
head­
dips)
was
significantly
higher
(
p
<
0.05)
in
the
aldrin­
treated
groups
when
compared
to
that
of
the
control
rats.
7­
12
Aldrin/
Dieldrin
 
February
2003
7.2.5
Developmental/
Reproductive
Toxicity
Oral
Exposure
Aldrin
In
a
reproduction
study
reported
by
Deichmann
et
al.
(
1971),
groups
of
beagles
were
administered
0.15
(
4
females)
or
0.3
(
4
males,
3
females)
mg/
kg
bw/
day
of
aldrin
(
purity
95%)
by
capsule,
5
days/
week
for
14
months.
Estrous
cycles
in
the
female
dogs
were
delayed
by
7
to
12
months,
and
2
of
the
4
females
administered
0.15
mg/
kg
bw/
day
failed
to
achieve
estrus
during
the
8­
month
period
following
cessation
of
aldrin
exposure.
However,
such
failure
was
not
observed
in
dogs
given
0.3
mg/
kg
bw/
day.
During
lactation,
the
viability
of
pups
from
dams
receiving
either
0.15
or
0.3
mg/
kg
bw/
day
was
decreased;
84,
75,
and
44%
of
pups
from
dams
ingesting
0,
0.15,
and
0.3
mg/
kg
bw/
day,
respectively,
survived
until
weaning.
The
reduced
pup
survival
may
have
been
due
to
a
prenatal
effect,
or
to
toxicity
associated
with
dieldrin
in
the
mothers'
milk.
Mammary
development
and
milk
production
also
appeared
to
be
severely
depressed.
Some
males
reportedly
exhibited
a
depressed
sexual
drive.

Dieldrin
Coulston
et
al.
(
1980)
studied
the
reproductive
effects
of
dieldrin
on
Long
Evans
rats.
Pregnant
rats
(
18
to
20/
dose)
were
administered
0
or
4
mg/
kg
bw
dieldrin
(
purity
not
reported)
by
gavage,
daily
from
day
15
of
gestation
through
postpartum
day
21.
The
treated
group
did
not
differ
from
the
control
group
when
examined
for
fecundity,
number
of
stillbirths,
perinatal
mortality,
or
total
litter
weights.
Pup
malformations
were
not
observed
in
either
group.

Harr
et
al.
(
1970)
fed
dieldrin
(
purity
not
specified)
to
28
day­
old
OSU­
Wistar
rats
(
20/
sex/
group)
until
they
were
mated
at
146
days
of
age;
dietary
concentrations
were
0,
0.08,
0.16,
0.31,
0.63,
1.25,
2.5,
5,
10,
20,
or
40
mg/
kg
(
0
and
approximately
0.004,
0.008,
0.016,
0.032,
0.063,
0.125,
0.25,
0.5,
1,
or
2
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959).
Mortality
was
observed
in
dams
exposed
to
1
or
2
mg/
kg
bw/
day,
and
fertility
and
litter
size
were
reduced
in
several
groups
without
demonstrating
a
clear
dose­
response
relationship.
At
weaning,
no
pups
survived
in
the
1
and
2
mg/
kg
bw/
day
groups,
and
the
number
of
survivors
was
substantially
reduced
at
doses
down
to
0.125
mg/
kg
bw/
day.
At
these
doses,
pups
died
in
convulsions
(
43%)
or
starved
(
57%),
the
latter
occurring
because
both
dams
and
pups
were
too
hyperesthetic
to
permit
adequate
nursing.
Neural
lesions
(
e.
g.,
cerebral
edema
and
hydrocephalus)
were
noted
in
pups
of
the
0.004
mg/
kg
bw/
day
group
(
but
evidently
not
in
those
of
higher­
dose
groups),
and
hepatic
lesions
were
observed
in
rats
exposed
to
$
0.016
mg
dieldrin/
kg
bw/
day.
This
study
has
been
considered
somewhat
limited
by
the
lack
of
statistical
analysis
and
by
the
uncertain
affect
on
outcome
that
may
have
resulted
from
the
use
of
a
semisynthetic
diet
(
ATSDR,
2000).

Dieldrin
(
87%
pure)
was
not
found
to
be
teratogenic
in
pregnant
CD
rats
(
9
to
25/
group)
and
CD­
1
mice
(
6
to
16/
group)
that
were
administered
doses
in
peanut
oil
of
0,
1.5,
3.0,
or
7­
13
Aldrin/
Dieldrin
 
February
2003
6.0
mg/
kg
bw/
day
by
gastric
intubation
on
days
7
through
16
of
gestation
(
Chernoff
et
al.,
1975).
Fetal
toxicity
was
reported
in
the
mice,
as
indicated
by
a
significant
decrease
in
the
numbers
of
caudal
ossification
centers
at
the
6.0
mg/
kg
bw/
day
dose
level,
and
a
significant
increase
in
the
number
of
supernumerary
ribs
in
one
study
group
at
both
the
3.0
and
6.0
mg/
kg
bw/
day
doses.
In
the
second
study
group,
the
increase
was
significant
only
at
the
3.0
mg/
kg
bw/
day
group.
In
contrast
to
these
results
in
mice,
exposed
rat
fetuses
evidenced
no
differences
from
controls
in
body
weight,
mortality,
or
the
occurrence
of
anomalies.
Maternal
toxicity
in
the
high­
dose
rats
was
indicated
by
a
4l%
mortality
and
a
significant
decrease
in
weight
gain;
similarly,
mice
receiving
6.0
mg/
kg
bw/
day
showed
a
significant
decrease
in
maternal
weight
gain.
A
significant
increase
in
liver­
to­
body
weight
ratio
in
one
group
of
maternal
mice
was
reported
at
both
the
3.0
and
6.0
mg/
kg
bw/
day
doses.
Thus,
any
evidence
of
dieldrin's
potential
teratogenicity
was
accompanied
by
concomitant
maternal
toxicity.

CFW
Swiss
mice
(
100/
sex)
fed
5
mg
dieldrin/
kg
diet
(
purity
not
reported;
approximately
equal
to
0.75
mg/
kg
bw/
day
based
on
Lehman,
1959)
for
30
days
prior
to
mating,
and
then
for
90
days
thereafter,
experienced
no
adverse
effects
on
fertility,
fecundity,
length
of
gestation
period,
size
of
first
litters,
or
numbers
of
young
produced
per
day
(
Good
and
Ware,
1969).
The
only
adverse
reproductive
effect
observed
in
this
study
was
a
slight
decrease
(
6%)
in
mean
size
of
all
litters
combined.

Virgo
and
Bellward
(
1975)
fed
dieldrin
(
purity
not
reported)
to
uniparous
female
Swiss­
Vancouver
mice
(
18
to
19/
group)
at
dietary
concentrations
of
0,
2.5,
5,
10,
15,
20,
or
25
mg/
kg
(
0
and
approximately
0.375,
0.75,
1.5,
2.25,
3.0,
or
3.75
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
for
a
period
extending
from
4
weeks
prior
to
their
second
mating
through
postpartum
day
28.
Males
were
exposed
only
during
the
2­
week
mating
period.
Significant
preparturition
mortality
was
observed
in
3.0
and
3.75
mg/
kg
bw/
day
females
(
89
and
56%,
respectively),
while
fertility
was
decreased
in
the
1.5
and
2.25
mg/
kg
bw/
day
groups.
Estrus
and
gestation
period
were
unaffected
by
the
treatment,
but
litter
size
was
reduced
at
3.75
mg/
kg
bw/
day.
Pre­
weaning
pup
mortality
was
increased
from
31%
in
control
animals
to
47,
80,
or
100%
in
the
0.375,
0.75,
or
1.5
and
higher
mg/
kg
bw/
day
groups,
respectively.
Hyperactivity
was
exhibited
by
dams
exposed
to
1.5
or
more
mg/
kg
bw/
day,
which
was
a
contributing
factor
to
high
pup
mortality.
Some
higher­
dose
dams
violently
shook
their
pups,
ultimately
killing
them,
and
others
neglected
their
litters.
No
gross
abnormalities
were
observed
in
pups
from
any
dose
group.

In
a
subsequent
cross­
fostering
study,
Virgo
and
Bellward
(
1977)
fed
primiparous
female
Swiss­
Vancouver
mice
(
number/
group
not
reported
in
citing
references)
diets
containing
dieldrin
(
purity
not
reported)
at
concentrations
of
0,
5,
10,
or
15
mg/
kg
(
0
and
approximately
0.75,
1.5,
or
2.25
mg/
kg
bw/
day
based
on
Lehman,
1959)
for
4
weeks
prior
to
mating.
Nursing
was
reduced
in
dams
exposed
to
the
two
highest
doses
of
dieldrin,
although
serum
progesterone
levels,
milk
production,
and
the
dams'
tendencies
to
build
nests
or
retrieve
pups
were
not
adversely
affected.
When
foster
dams
not
exposed
to
dieldrin
nursed
pups
from
the
1.5
mg/
kg
bw/
day
group,
all
died
within
4
days;
the
foster
dams'
own
pups
evidenced
very
low
mortality
and
survived
until
weaning.
Similar
results
were
also
reported
for
pups
from
the
0.75
mg/
kg
bw/
day
group.
7­
14
Aldrin/
Dieldrin
 
February
2003
Aldrin/
Dieldrin
Treon
et
al.
(
1954)
reported
increased
mortality
during
the
first
5
days
of
life
in
offspring
from
the
first
mating
of
a
three­
generation
reproduction
study,
in
which
rats
were
exposed
to
0.275
mg/
kg
bw/
day
of
either
aldrin
or
dieldrin
(
purity
not
reported).
Reduced
fertility
during
the
parental
generation's
first
mating
was
reported
at
doses
of
aldrin
and
dieldrin
as
low
as
1.38
and
0.275
mg/
kg
bw/
day,
respectively.
Subsequent
parental
matings
did
not
demonstrate
reproductive
effects
in
the
aldrin­
exposed
groups,
while
fertility
effects
in
the
dieldrin­
exposed
groups
failed
to
exhibit
consistent
dose­
related
responses.
During
matings
of
the
offspring,
reductions
in
fertility
were
not
observed
at
the
0.275
mg/
kg
bw/
day
doses,
but
could
not
be
adequately
assessed
at
higher
doses
due
to
limited
numbers
of
offspring
surviving
to
be
mated.

In
a
three­
generation
study
by
Treon
and
Cleveland
(
1955),
groups
of
Carworth
rats
(
number/
group
not
reported)
were
fed
aldrin
or
dieldrin
(
95%
purity)
at
concentrations
of
0,
2.5,
12.5,
or
25.0
ppm
(
doses
of
0
and
approximately
0.125,
0.625,
or
1.25
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959).
Two
litters/
generation
were
produced.
No
reductions
in
the
numbers
of
live
pups/
litter
or
pup
weights
were
evident
in
dams
fed
any
dose
of
either
chemical.
However,
viability
of
the
offspring
during
lactation
was
markedly
decreased
in
the
0.625
and
1.25
mg/
kg
bw/
day
groups
for
both
chemicals,
and
slightly­
to­
moderately
decreased
in
the
low­
dose
groups.
Pregnancy
rates
were
reportedly
initially
reduced
at
the
mid
and
high
doses
of
aldrin,
and
at
all
three
doses
of
dieldrin;
this
effect,
however,
gradually
disappeared
over
successive
generations.

In
a
study
that
examined
two
litters/
generation
over
six
generations,
Keplinger
et
al.
(
1970)
fed
Swiss
white
mice
(
4M
to
14F/
group)
diets
containing
aldrin
(
purity
not
reported)
at
concentrations
of
0,
3,
5,
10,
or
25
mg/
kg
(
0
and
approximately
0.45,
0.75,
1.5,
or
3.75
mg/
kg
bw/
day,
respectively,
according
to
Lehman,
1959).
The
3.75
mg/
kg
bw/
day
group
was
discontinued
due
to
excessive
litter
mortality
in
the
few
dams
reaching
gestation.
Otherwise,
the
most
pronounced
effect
reported
was
a
reduction
in
suckling
pup
survival
at
1.5
mg/
kg
bw/
day,
and
to
a
lesser
degree
at
0.75
mg/
kg
bw/
day.
Similarly,
groups
of
mice
were
fed
diets
containing
dieldrin
at
concentrations
of
0,
3,
10,
or
25
mg/
kg
(
0
and
approximately
0.45,
1.5,
or
3.75
mg/
kg
bw/
day).
As
with
aldrin,
the
high
dieldrin
dose
was
soon
discontinued
for
reasons
of
excessive
litter
toxicity,
and
the
1.5
mg/
kg
bw/
day
dose
was
discontinued
after
the
first
generation
because
of
low
pup
survival.
At
the
remaining
0.45
mg/
kg
bw/
day
dose,
no
effects
on
fertility,
viability,
or
gestation
were
noted.
Although
a
decrease
in
suckling
pup
survival
was
observed
in
the
F
2b
litters,
a
similar
decrease
also
occurred
in
one
of
the
six
control
groups.

Ottolenghi
et
al.
(
1974)
exposed
pregnant
CD­
1
mice
(
10/
group)
and
Syrian
golden
hamsters
(
41
to
43/
group)
to
high
(
one
half
the
oral
LD
50),
single
oral
doses
of
either
aldrin
or
dieldrin
(>
99%
purity)
in
corn
oil.
Negative
control
groups
consisted
of
untreated
and
corn
oildosed
animals.
Mice
were
exposed
on
gestation
day
9
to
aldrin
at
25
mg/
kg
bw
or
dieldrin
at
15
mg/
kg
bw;
hamsters
on
either
gestation
day
7,
8,
or
9
to
aldrin
at
50
mg/
kg
bw,
or
dieldrin
at
30
mg/
kg
bw.
In
mice,
the
aldrin
treatment
did
not
affect
fetal
survival
or
weight,
but
significantly
increased
the
incidence
of
abnormalities
such
as
webbed
feet,
cleft
palate,
and
open
eyes
(
33%
of
the
live
fetuses
had
malformations).
In
hamsters,
aldrin
treatment
did
cause
a
reduction
in
fetal
7­
15
Aldrin/
Dieldrin
 
February
2003
survival
and
weight,
as
well
as
a
significant
increase
in
the
incidence
of
the
same
types
of
abnormalities
that
were
observed
in
mice;
these
effects
were
less
pronounced
when
treatment
was
on
gestation
day
9,
rather
than
on
days
7
or
8.
In
mice,
dieldrin
produced
the
same
types
of
abnormalities
(
in
17%
of
the
live
fetuses)
as
seen
with
aldrin
and
the
effects
in
hamsters
were
also
similar
to
those
described
for
aldrin
with
respect
to
fetal
toxicity,
malformation
types,
and
degree
of
severity
according
to
day
of
treatment.

No
apparent
effects
on
the
fertility
or
pregnancy
rates
were
evident
in
groups
of
mongrel
dogs
(
2/
group,
at
least
1/
each
sex)
receiving
0,
0.2,
0.6,
or
2.0
mg/
kg
bw/
day
of
either
aldrin
or
dieldrin
(
purity
99%)
in
medicated
meatballs
for
1
year
(
Kitselman,
1953).
However,
the
majority
of
apparently
healthy
pups
that
were
delivered
from
dams
in
all
dose
groups
of
aldrin,
and
from
the
mid­
and
high­
dose
groups
of
dieldrin,
died
within
3
days
postpartum
and
evidenced
degenerative
liver
and
renal
tubule
changes
upon
histopathological
examination.
It
should
be
noted
that
this
study
had
several
limitations
with
respect
to
size
and
design
parameters.

In
dominant
lethal
studies,
Epstein
et
al.
(
1972)
and
Dean
et
al.
(
1975)
reported
no
unequivocal
adverse
effects
on
reproduction
subsequent
to
acute
exposure
of
male
mice
to
aldrin
at
doses
up
to
1
mg/
kg
bw/
day
for
a
period
of
5
days,
or
to
single
oral
doses
of
dieldrin
ranging
from
12.5
to
50
mg/
kg
bw.

Inhalation
Exposure
No
studies
were
obtained
that
investigated
the
developmental
or
reproductive
effects
of
aldrin
or
dieldrin
in
animals
following
inhalation
exposure.

Dermal
Exposure
No
studies
were
obtained
that
investigated
the
developmental
or
reproductive
effects
of
aldrin
or
dieldrin
in
animals
following
dermal
exposure.

Other
Routes
of
Exposure
Castro
et
al.
(
1992)
reported
the
effects
of
prenatal
exposure
to
aldrin
on
the
physical
and
behavioral
developments
of
rats
(
1
to
21
day­
old
and
90
day­
old
groups).
Pregnant
female
rats
(
10
to
20/
group)
were
subcutaneously
injected
with
either
aldrin
(
1.0
mg/
kg
bw)
or
its
vehicle
(
0.9%
sodium
chloride
plus
Tween
80),
from
day
1
of
pregnancy
until
delivery.
Pups
from
the
aldrin
group
evidenced
a
decreased
median
effective
time
(
TE
50)
for
incisor
teeth
eruption,
and
an
increased
TE
50
for
testes
descent;
other
parameters
indicative
of
physical
development,
such
as
pinna
detachment,
development
of
fur
and
ears,
and
eye
opening,
were
not
altered.
No
changes
in
body
weight
were
observed
between
control
and
aldrin
treated
rats
on
the
day
of
birth,
at
weaning,
or
at
90
days.
Prenatal
exposure
to
aldrin
produced
no
changes
in
aldrin
or
dieldrin
levels
in
serum,
or
in
cellular
and
structural
organization
of
cerebral
cortex
neurons,
when
tested
at
90
days.
7­
16
Aldrin/
Dieldrin
 
February
2003
Johns
et
al.
(
1998)
reported
no
significant
differences
in
birth
weight,
sex
ratio,
day
of
eye
opening,
or
weight
gain
between
the
pups
of
control
and
dieldrin­
treated
female
rats,
which
had
been
intraperitoneally
injected
daily
from
E12
to
E16
(
embryonic
days
12
to
16)
with
0,
5,
or
10
mg/
kg
bw
of
dieldrin.

7.2.6
Chronic
Toxicity
Oral
Exposure
Aldrin
Treon
and
Cleveland
(
1955)
administered
aldrin
in
the
diet
to
40
Carworth
rats/
sex
at
concentrations
of
2.5,
12.5,
or
25
ppm
(
approximate
doses
of
0.125,
0.65,
or
1.25
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
for
a
period
of
2
years.
Forty
animals/
sex
served
as
controls.
Mortality
of
the
treated
rats
was
greater
than
that
of
controls,
with
50%
surviving
in
the
2.5
and
12.5
ppm
groups
and
40%
surviving
in
the
25
ppm
group
at
the
end
of
the
experiment.

Fitzhugh
et
al.
(
1964)
fed
groups
(
12/
sex/
group)
of
Osborne­
Mendel
rats
aldrin
(
purity
99%)
in
the
diet
at
concentrations
of
0.5,
2,
10,
50,
100,
or
150
ppm
(
approximated
doses
of
0.025,
0.1,
0.5,
2.5,
5.0,
and
7.5
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
for
2
years.
A
dose­
related
increase
in
mortality
was
observed
at
dietary
levels
of
50
ppm
or
greater.
In
addition,
significant
(
p
#
0.05)
dose­
related
increases
in
relative
liver
weights
were
observed.
Histopathological
changes
observed
in
the
livers
of
aldrin­
treated
rats
were
referred
to
as
primarily
the
characteristic
"
chlorinated
insecticide"
lesions
that
occur
only
in
rodents.
These
lesions
consist
of
enlarged
centrilobular
hepatic
cells,
with
increased
cytoplasmic
oxyphilia,
and
peripheral
migration
of
basophilic
granules.
The
incidence
and
severity
of
these
nonneoplastic
histologic
changes
increased
with
increasing
dietary
aldrin
level.
In
rats
ingesting
amounts
of
aldrin
at
50
ppm
or
higher,
distended
and
hemorrhagic
urinary
bladders,
enlarged
livers,
and
increased
incidences
of
nephritis
were
reported.
The
apparent
LOAEL
for
this
study
was
0.5
ppm
(
0.025
mg/
kg
bw/
day),
while
a
NOAEL
was
not
established.

Deichmann
et
al.
(
1970)
fed
groups
of
Osborne­
Mendel
rats
(
50/
sex/
dose)
aldrin
(
technical
grade,
95%
pure)
for
31
months
at
concentrations
of
either
20,
30,
or
50
ppm
(
1,
1.5,
or
2.5
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959).
Groups
of
100
rats/
sex
served
as
controls.
Survival
and
body
weight
gains
were
comparable
between
the
treated
and
the
control
groups,
but
treated
animals
exhibited
tremors
and
clonic
convulsions.
Liver­
to­
body
weight
ratios
were
increased
in
males
fed
30
or
50
ppm
aldrin.
Moderate
increases
(
not
dose­
related)
in
the
incidences
of
hepatic
centrilobular
cloudy
swelling
and
necrosis
were
observed
in
all
aldrintreated
male
and
female
rats,
but
not
in
the
controls.
A
LOAEL
of
20
ppm
(
1
mg/
kg
bw/
day)
was
established
by
this
study,
but
a
NOAEL
was
not
determined.

Groups
of
Osborne­
Mendel
rats
(
50/
sex/
group)
were
exposed
to
30
or
60
ppm
of
aldrin
(
95%
purity)
in
the
diet
(
approximate
doses
of
1.5
or
3.0
mg/
kg
bw/
day,
based
on
Lehman,
1959)
for
74/
80
weeks
(
M/
F),
followed
by
32
to
38
weeks
of
observation
(
NCI,
1978).
Pooled
controls
(
58M/
60F
from
similar
bioassays,
plus
10M/
10F
concurrent
controls)
were
used
for
statistical
7­
17
Aldrin/
Dieldrin
 
February
2003
evaluations.
While
no
significant
effects
of
aldrin
exposure
on
mortality
were
observed,
mean
body
weight
gains
during
the
second
year
were
lower
than
control
values.
Signs
typical
of
organochlorine
intoxication
(
hyperexcitability,
tremors,
convulsions),
with
frequency
and
severity
increasing,
especially
during
the
second
year.
Routine
gross
and/
or
microscopic
evaluation
revealed
no
adverse,
non­
neoplastic
respiratory,
cardiovascular,
gastrointestinal,
musculoskeletal,
hepatic,
renal,
endocrine,
dermal,
or
ocular
effects
resulting
from
exposure
to
aldrin.

Aldrin
(
technical
grade,
95%
pure)
was
administered
in
the
diet
for
80
weeks
(
followed
by
10
to
13
weeks
of
observation)
at
concentrations
of
4
or
8
ppm
(
approximate
doses
of
0.6
or
1.2
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
to
groups
of
50
male
mice,
and
at
concentrations
of
3
or
6
ppm
(
approximate
doses
of
0.45
or
0.90
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
to
groups
of
50
female
mice
(
NCI,
1978).
Pooled
controls
(
92M/
79F
from
similar
bioassays,
plus
20M/
10F
concurrent
controls)
were
used
for
statistical
evaluations.
In
a
trend
test,
a
significant
(
p
=
0.037),
dose­
dependent
increase
in
mortality
was
observed
in
females;
a
similar
effect
was
not
observed
in
males.
Hyperexcitability
was
observed
in
all
exposed
groups,
with
frequency
and
severity
increasing
during
the
second
year.
Mean
body
weight
was
unaffected
during
the
first
year,
but
somewhat
lower
than
control
values
during
the
second
year.
Routine
gross
and/
or
microscopic
evaluation
revealed
no
adverse,
non­
neoplastic
respiratory,
cardiovascular,
gastrointestinal,
musculoskeletal,
hepatic,
renal,
endocrine,
dermal,
or
ocular
effects
resulting
from
exposure
to
aldrin.
A
NOAEL
was
not
established
because
of
toxicity
at
3
ppm
(
0.45
mg/
kg
bw/
day),
the
lowest
dose
tested.

Kitselman
and
Borgmann
(
1952)
fed
groups
of
7
mongrel
dogs
of
both
sexes
(
number/
sex
not
specified)
either
0.2,
0.6,
or
2
mg/
kg
bw/
day
of
aldrin
in
medicated
meatballs,
for
up
to
228
days.
The
test
material
was
reported
to
have
been
99%
pure.
Dogs
that
were
administered
the
2
mg/
kg
bw/
day
dose
exhibited
marked
body
weight
loss,
and
they
all
died
between
days
60
and
90.
No
treatment­
related
effects
were
observed
in
dogs
receiving
the
0.2
mg/
kg
bw/
day
dose
for
190
days,
or
in
those
administered
the
0.6
mg/
kg
bw/
day
dose
for
228
days.
Based
on
body
weight
loss,
0.6
mg/
kg
bw/
day
and
2
mg/
kg
bw/
day
were
considered
to
be
the
NOAEL
and
LOAEL,
respectively,
for
this
study.

In
a
long­
term
feeding
study
by
Treon
and
Cleveland
(
1955),
beagles
(
2/
sex/
dose)
fed
diets
containing
aldrin
(
purity
95%)
at
concentrations
of
1
or
3
ppm
(
approximate
doses
of
0.043
to
0.091
or
0.12
to
0.25
mg/
kg
bw/
day,
respectively,
as
reported
by
the
authors)
for
15.6
months,
gained
weight
at
rates
similar
to
control
dogs.
However,
at
3
ppm,
significant
(
p
<
0.05)
increases
in
absolute
and
relative
liver
weights
were
noted.
Histopathologic
changes,
such
as
fatty
degeneration
of
the
liver
and
vacuolation
of
renal
tubular
cells,
were
also
observed
in
both
sexes
at
the
3
ppm
level.
At
the
1
ppm
level,
females
exhibited
vacuolation
of
the
distal
renal
tubules.
The
LOAEL
for
this
study
was
1
ppm
(
0.043
to
0.091
mg/
kg
bw/
day),
while
a
NOAEL
was
not
established.

Fitzhugh
et
al.
(
1964)
administered
0.2,
0.5,
1,
2,
or
5
mg/
kg
bw/
day
aldrin
(
purity
99%)
to
12
mongrel
dogs
(
sexes
combined),
6
days/
week,
for
periods
of
up
to
25
months.
Each
group
consisted
of
one
dog/
sex
except
for
the
0.5
mg/
kg
bw/
day
group,
which
had
one
male
and
three
7­
18
Aldrin/
Dieldrin
 
February
2003
female
dogs.
All
dogs
receiving
1,
2,
or
5
mg/
kg
bw/
day
died
within
49
weeks;
the
first
death
occurred
on
day
22
in
a
female
administered
5
mg/
kg
bw/
day.
Prior
to
death,
the
animals
exhibited
body
weight
loss,
dehydration,
and
convulsions.
Slight
to
moderate
fatty
degeneration
was
noted
in
hepatic
and
renal
tubular
cells
and
reduced
numbers
of
mature
erythroid
cells
were
found
in
the
bone
marrow.
In
animals
receiving
0.5
mg/
kg
bw/
day,
clinical
signs
of
toxicity
were
limited
to
convulsions
in
one
male
dog
during
the
24th
month.
Dogs
in
the
0.2
mg/
kg
bw/
day
group
exhibited
no
adverse
effects.
The
NOAEL
in
this
study
thus
appears
to
be
0.2
mg/
kg
bw/
day,
based
on
the
absence
of
clinical
signs
of
toxicity,
body
weight
loss,
and
histopathological
changes.
However,
the
adequacy
of
this
study
for
establishing
a
reliable
NOAEL
is
limited
by
the
small
number
of
dogs
used.

Dieldrin
Groups
of
Osborne­
Mendel
rats,
12/
sex/
dose,
were
fed
0,
0.5,
2,
10,
50,
100,
or
150
ppm
dieldrin
(
recrystallized,
100%
active
ingredient)
in
their
diet
for
2
years
(
Fitzhugh
et
al.,
1964).
These
concentrations
correspond
to
doses
of
0
and
approximately
0.025,
0.1,
0.5,
2.5,
5.0,
or
7.5
mg/
kg
bw/
day,
respectively,
based
on
Lehman
(
1959).
Survival
was
markedly
decreased
at
levels
of
50
ppm
and
above.
Liver­
to­
body
weight
ratios
were
significantly
increased
at
all
treatment
levels,
with
females
showing
the
effect
beginning
at
0.5
ppm,
and
males
at
$
10
ppm.
Microscopic
lesions
were
described
as
being
characteristic
of
chlorinated
hydrocarbon
exposure.
These
changes
were
minimal
at
the
0.5
ppm
level.
Male
rats,
at
the
two
highest
dose
levels
(
100
and
150
ppm),
developed
hemorrhagic
and/
or
distended
urinary
bladders,
usually
associated
with
considerable
nephritis.
A
LOAEL
of
0.025
mg/
kg
bw/
day,
the
lowest
dose
tested,
was
identified
in
this
study.

Groups
of
Carworth
Farm
"
E"
strain
rats
(
25/
sex/
dose
level)
were
fed
dieldrin
(>
99%
purity)
in
the
diet
at
concentrations
of
0,
0.1,
1.0,
or
10.0
ppm
for
2
years.
These
doses
correspond
to
doses
of
0
and
approximately
0.005,
0.05,
or
0.5
mg/
kg
bw/
day,
respectively,
based
on
Lehman
(
1959).
At
7
months,
the
1
ppm
intake
level
was
equivalent
to
approximately
0.05
and
0.06
mg/
kg
bw/
day
for
males
and
females,
respectively.
No
effects
on
mortality,
body
weight,
food
intake,
hematology
or
blood,
and
urine
chemistries
were
reported.
At
the
10
ppm
level,
all
animals
became
irritable
after
8
to
13
weeks
of
treatment
and
developed
tremors
and
occasional
convulsions.
Liver
weights
and
liver­
to­
body
weight
ratios
were
significantly
increased
in
females
receiving
both
1.0
and
10
ppm.
Pathological
findings,
described
as
organochlorine­
insecticide
changes
of
the
liver,
were
found
in
one
male
and
six
females
at
the
10
ppm
level.
No
evidence
of
tumorigenesis
was
found
(
Walker
et.
al.,
1969).
Based
on
the
significantly
increased
liver
weight
and
relative
liver
weight
reported
for
female
rats,
this
study
establishes
a
NOAEL
and
a
LOAEL
of
0.005
and
0.05
mg/
kg
bw/
day,
respectively.

Walker
et
al.
(
1972)
administered
dieldrin
(>
99%
pure)
to
groups
of
CF1
mice
(
30/
sex/
dose)
in
the
diet
for
128
weeks
at
concentrations
of
1.25,
2.5,
5,
10,
or
20
ppm
(
approximate
doses
of
0.19,
0.38,
0.75,
1.5,
or
3
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959).
At
the
20
ppm
dose
level,
approximately
25%
of
the
males
and
nearly
50%
of
the
females
died
during
the
first
3
months
of
the
experiment.
Palpable
intra­
abdominal
masses
were
detected
after
40,
75,
or
100
weeks
in
the
10,
5,
and
2.5
ppm­
treated
groups,
respectively.
At
1.25
ppm,
7­
19
Aldrin/
Dieldrin
 
February
2003
liver
enlargement
was
not
palpable
and
morbidity
was
similar
to
that
of
controls.
A
NOAEL
cannot
confidently
be
established
from
this
study
because
clinical
chemistry
parameters
were
not
determined.

Groups
of
Osborne­
Mendel
rats
(
50/
sex/
group)
were
exposed
to
29
or
65
ppm
of
dieldrin
(
95%
purity)
in
the
diet
(
approximate
doses
of
1.45
or
3.25
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
for
80
weeks
followed
by
30
to
31
weeks
of
observation
for
the
low
dose,
or
for
59
weeks
followed
by
51
to
52
weeks
of
observation
for
the
high
dose
(
NCI,
1978).
Pooled
controls
(
58M/
60F
from
similar
bioassays,
plus
10M/
10F
concurrent
controls)
were
used
for
statistical
evaluations.
While
no
statistically
significant
end­
result
effects
of
dieldrin
exposure
on
mortality
were
observed,
it
was
perhaps
accelerated
in
treated
animals,
and
mean
body
weight
gains
during
the
second
year
were
lower
than
control
values.
Signs
typical
of
organochlorine
intoxication
(
hyperexcitability,
tremors,
convulsions)
were
evident,
with
frequency
and
severity
increasing,
especially
during
the
second
year
and
in
high­
dose
animals.
Routine
gross
and/
or
microscopic
evaluation
revealed
no
adverse,
non­
neoplastic
respiratory,
cardiovascular,
gastrointestinal,
musculoskeletal,
hepatic,
renal,
endocrine,
dermal,
or
ocular
effects
resulting
from
exposure
to
dieldrin.

In
a
related
study,
groups
of
Fischer
344
rats
(
24/
sex/
group)
were
exposed
to
2,
10,
or
50
ppm
of
dieldrin
("
purified
technical
grade")
in
the
diet
(
approximate
doses
of
0.1,
0.5,
or
2.5
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
for
104
to
105
weeks
(
NCI,
1978).
Body
weight
and
mortality
were
not
significantly
affected
by
dieldrin
exposure,
but
signs
typical
of
organochlorine
intoxication
(
hyperexcitability,
tremors,
convulsions)
were
noted
in
both
sexes
at
the
high
dose
after
80
weeks.
As
in
the
previously
discussed
study,
no
other
significant
adverse
systemic
effects
were
observed.

Dieldrin
(
95%
pure)
was
administered
in
the
diet
for
80
weeks
(
followed
by
10
to
13
weeks
of
observation)
at
concentrations
of
2.5
or
5
ppm
(
approximate
doses
of
0.375
or
0.75
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
to
groups
of
B6C3F
1
mice
(
50/
sex/
group)
(
NCI,
1978).
Pooled
controls
(
92M/
79F
from
similar
bioassays,
plus
20M/
10F
concurrent
controls)
were
used
for
statistical
evaluations.
Treatment
had
no
appreciable
effect
on
survival,
while
weight
gains
were
non­
significantly
lower
than
control
values
during
the
second
year.
Hyperexcitability,
hyperactivity,
fighting,
and
tremors
were
found
to
be
treatment­
related,
and
were
first
observed
in
males,
then
later
in
females.
Routine
gross
and/
or
microscopic
evaluation
revealed
no
adverse,
non­
neoplastic
respiratory,
cardiovascular,
gastrointestinal,
musculoskeletal,
hepatic,
renal,
endocrine,
dermal,
or
ocular
effects
resulting
from
exposure
to
aldrin.

Mongrel
dogs,
1/
sex/
dose
(
2/
sex
at
0.5
mg/
kg
bw/
day),
that
were
fed
dieldrin
(
recrystallized,
100%
active
ingredient)
at
dose
levels
of
0.2
to
10
mg/
kg
bw/
day,
6
days/
week
for
up
to
25
months,
showed
various
toxic
effects,
including
weight
loss
and
convulsions
at
dosages
of
0.5
mg/
kg
bw/
day
or
more.
Survival
was
inversely
proportional
to
dose
level.
No
toxic
effects,
gross
or
microscopic,
were
seen
at
a
dose
level
of
0.2
mg/
kg
bw/
day
(
Fitzhugh
et.
al.,
1964).
A
NOAEL
of
0.2
mg/
kg
bw/
day
appears
to
have
been
established
for
this
study,
but
its
reliability
is
substantially
limited
because
of
the
low
number
of
animals
studied.
7­
20
Aldrin/
Dieldrin
 
February
2003
Groups
of
beagle
dogs
(
5/
sex/
dose)
were
treated
daily
by
capsule
with
dieldrin
(>
99%
purity)
at
0.0,
0.005,
or
0.05
mg/
kg
in
olive
oil
for
2
years.
No
treatment­
related
effects
were
seen
in
general
health,
behavior,
body
weight,
or
urine
chemistry.
A
significant
increase
in
plasma
alkaline
phosphatase
activity
in
both
sexes
and
a
significant
decrease
in
serum
protein
concentration
in
males
receiving
the
high
dose
were
not
associated
with
any
clinical
or
pathological
change.
Liver
weight
and
liver­
to­
body
weight
ratios
were
significantly
increased
in
females
receiving
the
high
dose,
0.05
mg/
kg
bw/
day,
but
no
gross
or
microscopic
lesions
were
found.
There
was
no
evidence
of
tumorigenic
activity
(
Walker
et
al.,
1969).

Inhalation
Exposure
No
studies
were
obtained
that
examined
the
chronic
effects
of
aldrin
or
dieldrin
in
animals
after
chronic
inhalation
exposure.

Dermal
Exposure
There
is
one
available
study,
which
was
conducted
in
rabbits,
that
examined
the
chronic
effects
of
dermal
exposure
to
dieldrin.
Witherup
et
al.
(
1961)
reported
no
effects
on
lung
weight
or
pathology,
heart
weight
or
pathology,
liver
weight,
serum
proteins,
thymol
turbidity,
serum
alkaline
phosphatase,
or
pathology
in
a
chronic
study
in
which
rabbits
were
wrapped
with
material
containing
up
to
0.04%
dieldrin
for
up
to
52
weeks.
However,
this
study
is
limited
in
that
some
animals
were
treated
with
a
variety
of
drugs
to
control
"
extraneous"
diseases.

7.2.7
Carcinogenicity
Oral
Exposure
Aldrin
In
a
Food
and
Drug
Administration
(
FDA)
long­
term
carcinogenesis
bioassay,
Davis
and
Fitzhugh
(
1962)
exposed
a
group
of
215
C3HeB/
Fe
mice
(
numbers/
sex
were
not
provided,
but
the
group
was
reportedly
divided
approximately
equally
by
sex)
for
up
to
2
years
to
a
diet
containing
aldrin
(
purity
not
specified)
at
10
ppm,
constituting
a
dose
of
approximately
1.5
mg/
kg
bw/
day
using
the
conversion
factor
of
Lehman
(
1959).
The
average
long­
term
survival
rate
of
the
treated
group
was
approximately
2
months
less
than
that
of
the
controls,
although
these
rates
may
have
been
affected
by
intercurrent
diseases,
pneumonia
and
intestinal
parasitism.
Results,
reported
for
the
combined
sexes,
indicated
a
significant
(
p
<
0.001)
increase
in
the
number
of
treated
mice
with
hepatic
cell
adenomas
(
35/
215
or
23%)
when
compared
to
that
for
the
control
group
(
9/
217
or
7%).
These
hepatic
cell
adenomas
were
described
as
"
expanding
nodules
of
hepatic
parenchymal
tissue,
usually
with
altered
lobular
architecture,
and
morphologically
ranging
from
very
benign
lesions
to
borderline
carcinomas."
As
reported
by
Epstein
(
1975a),
an
independent
reevaluation
of
these
lesions
by
other
pathologists
concluded
that
most
were
liver
carcinomas.
Despite
the
short­
comings
of
poor
survival
rate,
lack
of
detailed
pathology,
loss
of
a
large
number
of
animals
to
the
study,
and
failure
to
report
the
results
7­
21
Aldrin/
Dieldrin
 
February
2003
separately
by
sex,
the
study
provided
evidence
for
aldrin's
hepatocarcinogenicity
to
this
strain
of
mouse.

In
an
FDA
follow­
up
to
the
previous
study,
aldrin
of
unspecified
purity
was
fed
to
groups
of
C3H
mice
(
100/
sex)
at
concentrations
of
0
or
10
ppm
(
approximately
1.5
mg/
kg
bw/
day
using
the
conversion
factor
of
Lehman,
1959)
for
up
to
2
years
(
Davis,
1965).
The
incidences
(
for
both
sexes
combined)
of
hepatic
hyperplasia
and
benign
hepatomas
in
the
treated
group
were
reported
to
be
approximately
double
those
of
the
controls,
whereas
the
incidence
of
hepatic
carcinomas
was
judged
to
be
about
the
same.
This
study
suffered
some
of
the
deficiencies
of
its
predecessor,
and
again
an
independent
review
concluded
that
most
of
the
hepatomas
were
actually
hepatocellular
carcinomas
(
Epstein,
1975a).
This
reevaluation
provided
incidences
of
hepatocellular
carcinomas
in
the
treated
vs.
control
group
of
82%
vs.
30%
for
males,
and
85%
vs.
4%
for
females,
both
significant
increases
at
p
<
0.05.

Song
and
Harville
(
1964)
fed
a
total
of
55
C3H
and
CBA
mice
15
ppm
of
aldrin
(
unspecified
purity)
for
an
unspecified
amount
of
time;
10
mice
served
as
controls.
In
a
companion
study,
mice
were
similarly
fed
dieldrin.
Seven
mice
treated
with
aldrin
or
dieldrin
were
reported
to
have
developed
liver
tumors
by
330
to
375
days;
however,
as
no
further
details
were
described,
this
report
provides
little
useful
information.

In
a
somewhat
more
recent
carcinogenicity
bioassay,
technical
grade
aldrin
(
95%
pure)
was
administered
in
the
diet
for
80
weeks
to
B6C3F
1
mice
(
50/
sex)
at
time­
weighted
averages
of
4
or
8
ppm
for
males,
and
at
3
or
6
ppm
for
females
(
NCI,
1978).
Based
on
Lehman
(
1959),
these
concentrations
approximate
doses
of
0.6
and
1.2
mg/
kg
bw/
day
(
males),
and
0.45
and
0.90
mg/
kg
bw/
day
(
females).
The
animals
were
observed
for
an
additional
10
to
13
weeks.
A
significant
(
p
#
0.001)
dose­
related
increase
in
the
incidence
of
hepatocellular
carcinomas
was
observed
in
male,
but
not
female,
mice
when
compared
to
matched
or
pooled
controls.
Tumor
incidences
were
3/
20,
17/
92,
16/
49,
and
25/
45
for
the
matched
control,
pooled
control,
low­
dose
male,
and
high­
dose
male
groups,
respectively.

When
compared
with
control
animals,
NCI
(
1978)
also
reported
increased
incidences
for
combined
follicular
cell
adenoma
and
carcinoma
of
the
thyroid
in
both
male
and
female
Osborne­
Mendel
rats.
Treated
animals
(
50/
sex)
were
fed
technical
grade
aldrin
(
95%
pure)
at
concentrations
of
30
or
60
ppm
(
1.5
and
3
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
for
74
or
80
weeks
(
males
or
females,
respectively),
then
observed
for
an
additional
37
to
38
or
32
to
33
weeks
(
males
or
females,
respectively).
The
combined
incidences
from
the
pooled
control,
low­
dose,
and
high­
dose
groups
were
respectively
4/
48,
14/
38,
and
8/
38
for
males,
and
3/
52,
10/
39,
and
7/
46
for
females.
Differences
were
significant
(
p
=
0.001)
for
the
low­
dose
groups,
but
not
for
the
high­
dose
groups.
A
significant
(
p
=
0.001)
increase
in
the
incidence
of
cortical
adenomas
of
the
adrenal
gland
were
also
observed
in
the
low­
dose
females,
but
this
was
not
considered
to
be
compound
related
by
the
authors.
Aldrin
produced
no
significant
effect
on
the
mortality
of
rats
of
either
sex.
Overall,
the
authors
concluded
that
none
of
the
observed
tumors
were
associated
with
treatment,
a
view
that
has
been
echoed
elsewhere
(
USEPA,
1993a).
However,
other
evaluations
of
the
report
have
concluded
that
the
occurrence
of
the
thyroid
and
7­
22
Aldrin/
Dieldrin
 
February
2003
adrenal
cortex
tumors
should
be
considered
suggestive
or
equivocal
evidence
of
aldrin's
potential
carcinogenicity
in
the
rat
(
Griesemer
and
Cueto,
1980;
Haseman
et
al.,
1987;
USEPA,
1987).

A
number
of
other
carcinogenicity
bioassays
utilizing
Carworth
rats
(
Treon
and
Cleveland,
1955),
Holtzman
rats
(
Song
and
Harville,
1964),
or
Osborne­
Mendel
rats
(
Deichmann
et
al.,
1967,
1970;
Deichmann,
1974)
failed
to
find
evidence
of
aldrin­
induced
tumors,
but
all
suffered
from
substantial
experimental
and/
or
reporting
deficiencies
that
resulted
in
their
being
judged
inadequate
as
tests
of
aldrin's
possible
carcinogenicity
(
USEPA,
1987,
1993a).

Dieldrin
In
an
FDA
long­
term
carcinogenicity
bioassay,
Davis
and
Fitzhugh
(
1962)
exposed
groups
of
approximately
218
C3HeB/
Fe
mice
(
numbers/
sex
not
specified,
other
than
that
they
were
approximately
equal)
for
up
to
2
years
to
dieldrin
of
unspecified
purity
at
concentrations
in
the
diet
of
either
0
or
10
ppm
(
the
latter
corresponding
to
a
dose
of
approximately
1.5
mg/
kg
bw/
day,
Lehman,
1959).
Although
compromised
by
poor
survival
rates,
loss
of
a
large
percentage
of
the
animals
to
the
study
and
failure
to
treat
the
data
separately
by
sex,
the
study
did
demonstrate
a
significantly
increased
incidence
of
hepatomas
in
the
treated
group
when
compared
with
the
controls
(
36/
148
or
24%
vs.
9/
134
or
7%).
In
a
subsequent
follow­
up
study
by
FDA,
groups
of
C3H
mice
(
100/
sex)
were
fed
either
0
or
10
ppm
(
0
or
approximately
1.5
mg/
kg
bw/
day)
of
dieldrin
(
purity
not
specified)
for
up
to
2
years
(
Davis,
1965).
This
study
suffered
much
the
same
limitations
as
its
predecessor,
but
again
demonstrated
a
significant
increase
in
the
incidence
of
benign
hepatomas
(
and
in
the
combined
incidence
of
benign
hepatomas
plus
hepatocellular
carcinomas)
in
the
dieldrin
group
relative
to
controls.
As
for
the
companion
aldrin
studies
discussed
previously,
a
subsequent
pathology
reevaluation
of
both
of
these
studies
concluded
that
most
of
the
hepatomas
were
in
fact
malignant
hepatocellular
carcinomas
(
Epstein,
1975a,
b).

As
noted
previously,
Song
and
Harville
(
1964)
fed
a
total
of
55
C3H
and
CBA
mice
15
ppm
of
dieldrin
(
unspecified
purity)
for
an
unspecified
amount
of
time;
10
mice
served
as
controls.
In
a
companion
study,
mice
were
similarly
fed
aldrin.
Seven
mice
treated
with
aldrin
or
dieldrin
were
reported
to
have
developed
liver
tumors
by
330
to
375
days;
however,
as
no
further
details
were
described,
this
report
provides
little
useful
information.

Epstein
(
1975a)
reviewed
and
provided
reevaluations
of
an
unpublished
study
by
MacDonald
et
al.
(
1972),
in
which
"
technical
grade"
dieldrin
was
fed
for
an
uncertain
period
of
time
to
groups
of
Swiss­
Webster
mice
(
100/
sex/
group)
at
dietary
concentrations
of
either
0,
3,
or
10
ppm
(
corresponding
to
approximate
doses
of
0,
0.45,
or
1.5
mg/
kg
bw/
day,
respectively,
Lehman,
1959).
The
authors
concluded
that
dieldrin
was
not
carcinogenic,
but
that
it
induced
various
nonneoplastic
lesions
of
the
liver,
including
a
dose­
dependent
increase
in
the
incidence
of
hepatic
nodules
(
0,
2.5,
and
48%
at
0,
3,
and
10
ppm,
respectively).
However,
a
reevaluation
of
some
of
the
histopathological
data
by
independent
pathologists
(
as
well
as
by
one
of
the
original
authors)
demonstrated
that
more
than
half
of
the
reexamined
livers
from
high­
dose
mice
contained
hepatocellular
carcinoma,
thus
confirming
dieldrin's
carcinogenicity
to
mice.
7­
23
Aldrin/
Dieldrin
 
February
2003
Walker
et
al.
(
1972)
conducted
a
number
of
studies
in
which
they
exposed
groups
of
CF
1
mice
(
29
to
200/
sex/
dose;
29
to
300
controls/
sex/
study)
for
2
to
132
weeks
to
dietary
concentrations
of
dieldrin
(>
99%
pure)
ranging
from
0.1
to
20
ppm
(
approximating
doses
of
0.015
to
3.0
mg/
kg
bw/
day,
Lehman,
1959).
Significant
dose­
related
increases
in
the
incidences
of
benign
and
total
liver
tumors
were
observed
beginning
at
concentrations
as
low
as
2.5
ppm,
while
the
incidence
of
malignant
liver
tumors
was
significantly
increased
at
concentrations
of
5,
10,
and
20
ppm.
Liver
tumors
were
also
demonstrated
to
occur
much
earlier
in
treated
than
in
control
mice.
In
one
of
the
studies,
dieldrin
also
induced
significant
increases
(
p
<
0.05)
in
the
incidences
of
lung,
lymphoid,
and
"
other"
tumors
in
female
mice.

In
another
study
using
CF
1
mice
(
Thorpe
and
Walker,
1973),
groups
(
30/
sex;
45
controls/
sex)
were
fed
dieldrin
(>
99%
pure)
in
the
diet
for
up
to
110
weeks
at
concentrations
of
0
or
10
ppm
(
an
approximate
dose
of
1.5
mg/
kg
bw/
day,
Lehman,
1959).
Again,
a
statistically
significant
(
p
<
0.01)
increase
in
malignant
liver
tumors
(
many
of
which
metastasized
to
the
lung)
and
a
shortened
latency
period
were
induced
by
dieldrin.

In
an
NCI
(
1978)
study,
B6C3F
1
mice
(
50/
sex/
dose)
were
fed
technical
grade
dieldrin
(>
96%
purity)
for
80
weeks
(
with
an
additional
10
to
13
weeks
of
observation)
at
time­
weighted
average
concentrations
of
2.5
or
5
ppm
(
equivalent
to
0.375
or
0.75
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959).
Matched
(
20
male,
10
female)
and
pooled
(
92
male,
91
female)
controls
received
no
dieldrin
(
or
other
test
chemical)
in
their
feed.
This
assay
was
considered
an
acceptable
test
for
carcinogenicity
based
on
achieving
a
maximum
tolerated
dose
without
excess
toxicity
or
mortality
(
USEPA,
1987).
When
compared
with
pooled
controls,
male
mice
evidenced
a
significant
(
p
=
0.02)
dose­
related
increase
in
the
incidence
of
hepatocellular
carcinomas,
as
well
as
a
significant
(
p
=
0.025)
increase
in
such
tumors
at
the
high
dose.

In
a
study
by
Tennekes
et
al.
(
1981,
1979),
groups
of
19
to
82
male
CF
1
mice
were
fed
dieldrin
(>
99%
pure)
at
concentrations
of
0
or
10
ppm
(
an
approximate
dose
of
1.5
mg/
kg
bw/
day,
Lehman,
1959)
for
up
to
110
weeks.
Two
types
of
diet
and
two
types
of
bedding
were
examined
as
part
of
the
study.
Dieldrin
treatment
was
reported
to
have
shortened
the
liver
tumor
latency
period,
increased
the
incidence
of
combined
liver
tumors
from
10
to
81%,
and
significantly
(
p
<
0.01)
increased
the
incidences
of
hepatocellular
carcinoma
(
from
1
to
39%)
and
lung
metastases
(
from
0
to
14%).

In
a
large
study
intended
to
investigate
dieldrin's
enhancing
affect
on
liver
tumor
formation
(
Tennekes
et
al.,
1982),
a
total
of
1,800
CF
1
mice
(
17
to
297/
sex/
dose)
were
fed
dieldrin
(>
99.9%
purity)
over
the
course
of
their
lifetimes
at
concentrations
of
0,
0.1,
1,
2.5,
5,
10,
or
20
ppm
(
doses
of
0
and
approximately
0.015,
0.15,
0.375,
0.75,
1.5,
or
3.0
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959).
In
both
sexes,
treatment
appeared
to
result
in
dose­
related
increases
in
the
incidences
of
both
combined
(
benign
plus
malignant)
and
malignant
liver
tumors
up
to
10
ppm;
somewhat
lower
incidences
at
20
ppm
were
speculated
to
result
from
significant
toxicity/
lethality
at
that
concentration.
Dieldrin
also
induced
a
dose­
dependent
reduction
in
tumor
latency
periods;
the
lowest
doses
associated
with
a
significant
(
p
<
0.05)
reduction
in
median
time­
to­
tumor
formation
were
0.1
and
1.0
ppm
for
females
and
males,
respectively.
The
lack
of
a
linear
relationship
between
daily
exposure
level
and
median
time­
to­
tumor
formation
or
7­
24
Aldrin/
Dieldrin
 
February
2003
median
total
dose
led
the
authors
to
speculate
that
dieldrin
may
affect
tumor
promotion
rather
than
initiation.

Meierhenry
et
al.
(
1983)
exposed
groups
of
male
C3H/
He,
B6C3F
1,
and
C
57
BL/
6J
mice
(
50
to
71/
strain;
50
to
76
controls/
strain)
for
85
weeks
(
followed
by
47
weeks
of
observation)
to
dieldrin
(>
99%
purity)
at
a
dietary
concentration
of
10
ppm
(
an
approximate
dose
of
1.5
mg/
kg
bw/
day,
based
on
Lehman,
1959).
Dieldrin
induced
significant
(
p
<
0.05)
increases
in
the
incidences
of
hepatocellular
carcinomas
relative
to
controls
in
all
three
strains
of
mice.

Osborne­
Mendel
rats
treated
with
dieldrin
(>
96%
purity)
at
time­
weighted
average
concentrations
of
29
or
65
ppm
in
the
diet
(
approximate
doses
of
1.45
or
3.25
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
for
80
weeks,
and
then
observed
for
an
additional
30
to
31
weeks,
did
not
show
any
treatment­
related
increase
in
tumors
(
NCI,
1978).
A
second
NCI
(
1978)
study
that
exposed
groups
(
24/
sex)
of
Fischer
rats
to
dieldrin
(
technical
grade,
purified)
for
104
to
105
weeks
at
dietary
concentrations
of
0,
2,
10,
or
50
ppm
(
doses
of
0
and
approximately
0.1,
0.5,
or
2.5
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
produced
similarly
negative
tumorigenic
results.
Both
of
these
bioassays
were
judged
to
be
adequate
tests
for
carcinogenicity
(
USEPA,
1987).

As
evaluated
by
USEPA
(
1987),
one
other
minimally
acceptable
study
(
Deichmann
et
al.,
1970)
and
four
inadequate
studies
(
Treon
and
Cleveland,
1955;
Fitzhugh
et
al.,
1964;
Song
and
Harville,
1964;
Walker
et
al.,
1969,
which
was
reevaluated
by
Stevenson
et
al.,
1976)
collectively
exposed
several
strains
of
rats
(
Carworth,
Osborne­
Mendel,
or
Holtzman)
to
dietary
concentrations
of
dieldrin
(
varying
purities)
ranging
from
0.1
to
285
ppm
(
approximate
doses
of
0.005
to
14.25
mg/
kg
bw/
day,
respectively,
based
on
Lehman,
1959)
for
periods
of
1
to
2
years.
Although
all
of
these
studies
failed
to
demonstrate
any
evidence
for
dieldrin's
potential
carcinogenicity,
all
suffered
from
one
or
more
serious
deficiencies
(
e.
g.,
too
few
animals,
excessive
mortality,
inadequate
duration,
data
missing
or
inadequately
reported,
etc.).
Additionally,
several
dieldrin
bioassays
involving
dogs
or
monkeys
were
evaluated
by
the
USEPA
(
1987)
as
being
inadequate
or
unacceptable
tests
of
potential
carcinogenicity
due
to
serious
limitations.

Inhalation
Exposure
No
studies
were
obtained
that
examined
the
carcinogenicity
of
either
aldrin
or
dieldrin
in
animals
after
inhalation
exposure.

Dermal
Exposure
No
studies
were
obtained
that
examined
the
carcinogenicity
of
either
aldrin
or
dieldrin
in
animals
after
dermal
exposure.
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2003
7.3
Other
Key
Data
7.3.1
Mutagenicity/
Genotoxicity
Effects
Aldrin
In
bacterial
reverse
mutation
assays
that
were
conducted
by
several
investigators,
aldrin
was
not
mutagenic
to
Salmonella
typhimurium
(
Simmon
and
Kauhanen,
1978;
Cotruvo
et
al.
1977;
Simmon
et
al.,
1977;
Probst
et
al.,
1981;
Nishimura
et
al.,
1982)
or
E.
coli
(
Ashwood­
Smith
et
al.,
1972;
Probst
et
al.,
1981),
nor
was
it
found
to
induce
plasmid
DNA
breakage
in
E.
coli,
although
it
was
tested
only
in
the
absence
of
S9
metabolic
activation
(
Griffin
and
Hill,
1978).

Simmon
and
Kauhanen
(
1978)
reported
that
aldrin,
at
concentrations
of
10
to
5,000
:
g/
plate,
did
not
cause
gene
conversion
in
Saccharomyces
cerevisiae,
either
in
the
presence
or
absence
of
exogenous
metabolic
activation
provided
by
Aroclor­
induced
rat
liver
microsomes.
It
has,
however,
been
reported
to
induce
reverse
mutation
in
the
same
organism
(
Guerzoni
et
al.,
1976).

Several
doses
of
aldrin
were
tested
in
a
mouse
dominant
lethal
assay
conducted
by
Epstein
et
al.
(
1972),
and
although
some
reductions
in
the
level
of
implantation
were
demonstrated,
they
were
judged
to
be
statistically
nonsignificant.
Negative
results
have
also
been
reported
for
its
induction
of
sex­
linked
recessive
lethal
mutation
in
Drosophila
melanogaster
(
Benes
and
Sram,
1969).

Georgian
(
1975)
reported
that
aldrin
induced
chromosome
aberrations
in
human
lymphocytes
in
vitro
and
in
rat
and
mouse
bone
marrow
cells
in
vivo.
However,
the
evidence
for
an
in
vivo
clastogenic
response
is
somewhat
equivocal
because
the
observed
chromosomal
aberration
frequencies
increased
only
at
cytotoxic
levels.
Additionally,
chromosome
and
chromatid
gaps,
which
historically
have
been
considered
unreliable
indicators
of
significant
damage
to
genetic
material,
were
included
in
the
aberration
totals.
Therefore,
the
extent
of
the
more
meaningful,
non­
gap,
chromosomal
damage
cannot
be
ascertained.
Negative
results
have
also
been
reported
for
the
in
vivo
induction
of
micronuclei
in
mice
at
an
aldrin
dose
of
13
mg/
kg
bw
(
Rani
et
al.,
1980).

Dulout
et
al.
(
1985)
studied
the
incidences
of
sister
chromatid
exchanges
(
SCEs)
and
chromosome
aberrations
in
a
population
of
floriculturists
who
were
exposed
to
several
pesticides,
including
aldrin.
For
those
floriculturists
who
exhibited
clinical
symptoms
of
pesticide
exposure,
there
were
statistically
significant
increases
in
SCEs
when
compared
with
asymptomatic
floriculturists,
and
in
exchange­
type
chromosome
aberrations
when
compared
with
nonfloriculturists.
However,
interpretation
of
the
role
of
aldrin
in
these
findings
is
confounded
by
the
concomitant
exposure
to
other
organophosphorous,
carbamate,
and
organochlorine
pesticides.
Edwards
and
Priestly
(
1994)
reported
that
occupational
exposure
to
aldrin
did
not
alter
SCE
frequencies
in
lymphocytes
derived
from
workers
(
n
=
33)
recruited
from
two
south
Australian
suburban
pesticide
application
companies.
7­
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Aldrin/
Dieldrin
 
February
2003
Unscheduled
DNA
synthesis
(
UDS)
was
not
induced
when
primary
rat
hepatocytes
were
exposed
to
aldrin
at
concentrations
ranging
from
0.5
to
1,000
nmol/
mL
for
5
to
20
hours
(
Probst
et
al.,
1981),
and
most
likely
not
when
human
lymphocytes
were
exposed
to
concentrations
of
up
to
100
:
g/
mL
(
Rocchi
et
al.,
1980).
However,
Ahmed
et
al.
(
1977a)
reported
the
induction
of
UDS
in
transformed
human
cells
at
aldrin
concentrations
as
low
as
0.4
:
g/
mL,
and
Sina
et
al.
(
1983)
observed
DNA
strand
breaks
in
the
alkaline
elution/
rat
hepatocyte
assay
at
an
aldrin
concentration
of
110
:
g/
mL.

Dieldrin
Dieldrin
was
not
mutagenic
in
the
Salmonella/
microsome
test
(
Ames
test),
either
with
or
without
S­
9
mix
as
a
source
of
exogenous
metabolic
activation
(
McCann
et
al.,
1975).
Similarly,
nine
other
studies
have
collectively
reported
negative
responses
for
dieldrin
in
at
least
eight
different
Ames
tester
strains
of
S.
typhimurium,
both
with
and
without
exogenous
metabolic
activation
(
Anderson
and
Styles,
1978;
Bidwell
et
al.,
1975;
Glatt
et
al.,
1983;
Haworth
et
al.,
1983;
Marshall
et
al.,
1976;
Nishimura
et
al.,
1982;
Probst
et
al.,
1981;
Shirasu
et
al.,
1976;
Wade
et
al.,
1979).
Negative
responses
have
also
been
reported
for
dieldrin
in
E.
coli
using
both
a
reverse
mutation
assay
(
Ashwood­
Smith
et
al.,
1972;
Probst
et
al.,
1981)
and
two
forward
mutation
assay
systems
(
Gal
Rz2
and
streptomycin
resistance)
(
Fahrig,
1974).
Additionally,
Dean
et
al.
(
1975)
reported
negative
findings
in
a
host­
mediated
assay
(
microbial
cells
in
animal
hosts).
However,
in
one
contrary
study,
Majumdar
et
al.
(
1977)
reported
that
dieldrin
was
mutagenic
for
S.
typhimurium,
both
with
and
without
exogenous
metabolic
activation.

Dieldrin
produced
negative
responses
in
assays
for
forward
mutation
and
aneuploidy
induction
in
Aspergillus
nidulans
(
although
it
was
not
tested
with
exogenous
metabolic
activation)
(
Crebelli
et
al.,
1986),
gene
conversion
in
S.
cerevisiae,
and
reverse­
mutation
in
S.
marcesans
(
Fahrig,
1974),
in
vitro
DNA
strand
breaks
in
E.
coli
plasmids
or
in
animal
cell
alkaline
elution
assays
(
Swenberg
et
al.,
1976;
Swenberg,
1981),
UDS
in
rat
primary
hepatocytes
(
Probst
et
al.,
1981),
and
most
probably
for
UDS
in
human
lymphocytes
(
Rocchi
et
al.,
1980).
Dieldrin
also
failed
to
induce
cell
transformation
in
Syrian
hamster
embryo
cells
(
Mikalsen
and
Sanner,
1993).
However,
it
was
reported
to
induce
forward
mutation
in
Chinese
hamster
V79
cells
(
Ahmed
et
al.,
1977b),
as
well
as
UDS
in
transformed
human
cells
(
Ahmed
et
al.,
1977a).

In
a
dominant
lethal
study
that
orally
exposed
male
CF
1
mice
to
dieldrin,
mean
implantation
levels
(
versus
controls)
were
significantly
reduced
in
females
mated
with
males
receiving
12.5
mg/
kg
bw.
However,
in
a
subsequent
experiment,
mean
implantations
were
not
reduced,
or
even
significantly
increased,
in
females
mated
with
males
receiving
25
or
50
mg/
kg
bw
(
Dean
et
al.,
1975).
In
another
mouse
dominant
lethal
assay,
several
doses
of
dieldrin
up
to
26
mg/
kg
bw
were
found
to
be
without
mutagenic
effect
(
Epstein
et
al.,
1972).
Dieldrin
also
reportedly
did
not
induce
sex­
linked
recessive
mutation
in
D.
melanogaster
(
Benes
and
Sram,
1969).

Studies
have
demonstrated
that
dieldrin
can
cause
chromosomal
aberrations
in
mouse
bone
marrow
cells
following
in
vivo
exposure
(
Markaryan,
1966;
Dean
et
al.,
1975;
Majumdar
et
al.,
1976)
in
human
lymphoblastoid
cells
(
Trepanier
et
al.,
1977)
and
human
WI­
38
embryonic
7­
27
Aldrin/
Dieldrin
 
February
2003
lung
cells
(
Majumdar
et
al.,
1976)
after
in
vitro
exposure.
In
the
latter
case,
the
cytogenetic
effects
were
accompanied
by
significant
cytotoxicity,
as
there
was
evidence
of
cell
degeneration.
Dean
et
al.
(
1975)
failed
to
find
evidence
of
elevated
frequencies
of
chromosomal
aberrations
in
human
lymphocytes
after
in
vivo
exposure
to
undetermined
amounts
of
dieldrin.
SCEs,
but
not
chromosome
aberrations,
were
induced
by
dieldrin
in
CHO
cells,
both
in
the
presence
and
absence
of
S9
exogenous
metabolic
activation
(
Galloway
et
al.,
1987).
Aneuploidy
and/
or
nuclear
polyploidization
were
reportedly
induced
in
the
liver
of
CF
1
mice
treated
with
dieldrin
at
0.6
mg/
kg
bw/
day
(
van
Ravenzwaay
and
Kunz,
1988).
In
a
human
occupational
exposure
study,
Dean
et
al.
(
1975)
compared
the
frequencies
of
both
chromatid­
and
chromosome­
type
aberrations
in
lymphocytes
that
were
isolated
from
workers
exposed
to
dieldrin
in
a
manufacturing
facility
with
those
from
unexposed
control
subjects.
No
statistically
significant
differences
in
the
frequencies
were
observed.

With
respect
to
in
vivo
exposure,
currently
available
data
do
not
indicate
unequivocally
that
either
aldrin
or
dieldrin
directly
interacts
with
DNA
to
cause
mutations
in
either
the
germ
cells
or
the
somatic
cells
of
mammals.

7.3.2
Immunotoxicity
No
studies
were
obtained
that
examined
the
immunological
effects
of
aldrin
in
either
humans
or
animals,
and
only
limited
information
was
located
regarding
these
types
of
effects
in
humans
following
exposure
to
dieldrin.
A
case
report
was
located
concerning
a
man
who
developed
immunohemolytic
anemia
after
eating
fish
that
contained
high
levels
of
dieldrin
(
Hamilton
et
al.,
1978).
Testing
of
the
patient's
serum
revealed
a
positive
response
for
antibodies
to
dieldrin­
coated
red
blood
cells
(
RBCs).
Another
case
of
immunohemolytic
anemia
was
reported
in
a
man
who
had
had
multiple
exposures
to
dieldrin,
heptachlor,
and
toxaphene
while
spraying
cotton
fields
(
Muirhead
et
al.,
1959);
the
individual's
serum
was
found
to
contain
antibodies
to
RBCs
coated
with
either
dieldrin
or
heptachlor.
In
contrast,
volunteers
who
were
re­
exposed
to
fabric
that
contained
up
to
0.5%
dieldrin
2
weeks
after
an
initial
4­
day
exposure
did
not
reveal
any
evidence
of
sensitization
(
Suskind,
1959).

Immunosuppression
by
dieldrin
has
been
reported
in
a
number
of
studies
in
mice.
A
decrease
in
the
antigenic
response
to
the
mouse
hepatitis
virus
3
(
with
a
corresponding
increase
in
its
lethality)
was
observed
in
mice
given
a
single
oral
dose
of
dieldrin
($
18
mg/
kg)
(
Krzystyniak
et
al.,
1985).
Similarly,
an
increase
in
the
lethality
of
infections
with
the
malaria
parasite,
Plasmodium
berghei,
or
with
Leishmania
tropica
was
produced
in
mice
by
treatment
with
dieldrin
in
the
diet
at
concentrations
as
low
as
1
ppm
(
approximately
equivalent
to
0.15
mg/
kg
bw/
day
based
on
Lehman,
1959
for
10
weeks
[
Loose,
1982]).
In
addition,
decreased
tumor
cell
killing
ability
was
observed
in
mice
after
dieldrin
treatment
with
concentrations
as
low
as
1
ppm
(
approximately
equivalent
to
0.15
mg/
kg
bw/
day
based
on
Lehman,
1959)
for
3,
6,
or
18
weeks
(
Loose
et
al.,
1981).

Loose
et
al.
(
1981)
also
observed
a
decrease
in
antigen
processing
by
alveolar
macrophages
in
mice
following
administration
of
dieldrin
at
concentrations
as
low
as
0.5
ppm
(
approximately
equivalent
to
0.075
mg/
kg
bw/
day
based
on
Lehman,
1959)
for
2
weeks.
This
7­
28
Aldrin/
Dieldrin
 
February
2003
occurred
in
the
absence
of
observable
effects
on
macrophage
respiration,
phagocytic
activity
or
capacity,
or
microbial
activity.
In
addition,
macrophages
from
mice
exposed
for
10
weeks
to
dieldrin
in
the
diet
at
5
ppm
(
approximately
0.75
mg/
kg
bw/
day
based
on
Lehman,
1959)
were
found
to
produce
a
soluble
factor
that
induced
T­
lymphocyte
suppressor
cells,
suggesting
suppressed
immune
system
function
(
Loose,
1982).
In
another
limited
study,
lymphocyte
proliferation
appeared
inhibited
in
a
mixed
lymphocyte
reaction
test
in
which
splenic
cells
taken
from
mice
treated
twice
with
16.6
mg/
kg
bw
of
dieldrin
were
combined
with
stimulator
cells
taken
from
control
animals
(
Fournier
et
al.,
1988).

7.3.3
Hormonal
Disruption
Wade
et
al.
(
1997)
examined
hormone
levels
in
the
serum
and
uterine
tissues
of
young
female
Spargue­
Dawley
rats
after
intraperitoneal
exposure
to
3
mg/
kg­
day
dieldrin
from
days
18
to
21
after
birth.
As
compared
to
the
vehicle
treated
controls,
this
acute
exposure
to
dieldrin
produced
no
significant
effects
in
serum
thyroxine
levels,
or
in
uterine
tissue
levels
of
follicle
stimulating
hormone
(
FSH),
lutenizing
hormone
(
LH),
thyroid
stimulating
hormone,
prolactin,
or
growth
hormone.
Pituitary
weight
was
also
reported
to
be
unaltered
by
dieldrin
treatment.

In
an
in
vitro
study,
Brown
(
1998)
reported
that
very
low
dose
of
dieldrin
decreased
fetal
testicular
hormone
output.
Tissue
samples
from
6
human
male
fetuses,
terminated
after
12
to
19
weeks
of
gestation,
were
cultured
and
tested
for
the
production
of
testosterone
and
inhibin
after
exposure
to
dieldrin,
either
in
the
presence
or
absence
of
a
combination
of
FSH
and
LH
(
10
nM).
Diledrin
treatment
alone
did
not
reduce
hormone
secretions,
but
coadministration
of
FSH+
LH
and
dieldrin
(
10­
12
M)
significantly
reduced
(
p<
0.03)
testosterone
and
inhibin
B
levels
when
compared
to
control
levels.

7.3.4
Physiological
or
Mechanistic
Studies
One
mechanism
considered
as
a
possible
explanation
for
the
aldrin/
dieldrin­
induced
convulsions
and
tremors
observed
in
animals
and
humans
involves
the
effects
of
these
insecticides
on
the
GABA
(
gamma­
aminobutyric
acid)
receptor.
Several
lines
of
evidence
suggest
that
organochlorine
insecticides,
such
as
aldrin
and
dieldrin,
can
act
as
GABA
A
receptor
antagonists,
blocking
the
chloride
ion
channel
in
the
central
nervous
system.
Such
inhibition
of
the
chloride
ion
channel
could
be
a
significant
contributing
factor
to
convulsions
and
tremors
(
Klaassen,
1996).

Using
whole
cell
and
single­
channel
patch
clamp
techniques
Nagata
and
Narahashi
(
1995)
examined
dorsal
root
ganglion
(
DRG)
neurons,
isolated
from
the
lumbo­
dorsal
region
of
newborn
rats,
that
had
been
treated
with
dieldrin
(
0.0001
to
10
:
M).
They
reported
that
dieldrin
exposure
suppressed
the
GABA
A
receptor­
induced
chloride
currents
(
both
sensitive
and
lesssensitive
types)
in
a
time­
and
dose­
dependent
manner.
The
IC
50
values
were
estimated
as
being
3.7
nM
and
98
nM
for
the
sensitive
and
the
less­
sensitive
currents,
respectively.
Dieldrininduced
suppression
of
chloride
currents
were
directly
dependent
on
GABA
concentration
(
up
to
1000
:
M)
and
appeared
irreversible
as
the
current
did
not
recover
after
a
30­
minute
wash­
out
7­
29
Aldrin/
Dieldrin
 
February
2003
with
dieldrin­
free
solution.
This
suppression
of
chloride
currents
in
vitro
may
explain
the
in
vivo
hyperactivity
that
has
been
noted
in
animals
exposed
to
dieldrin.

Nagata
and
Narahashi
(
1994)
observed
that
dieldrin
(
1
:
M)
exhibited
dual
effects
on
chloride
currents
in
primary
cultures
of
DRG;
i.
e.,
initial
transient
enhancement
followed
by
suppression
after
repeated
co­
applications.
The
dieldrin­
caused
desensitization
and
suppression
of
chloride
current
occurred
at
an
EC
50
of
92
nM,
but
the
enhancement
of
chloride
current
needed
a
higher
EC
50
of
754
nM.
These
authors
also
reported
that
the
dieldrin­
induced
suppression
of
the
GABA­
mediated
chloride
current
was
non­
competitive
and
irreversible,
as
recovery
was
not
observed
after
prolonged
washing
of
the
neurons
with
dieldrin­
free
solution.

Nagata
et
al.
(
1994)
speculated
that
dieldrin
may
cause
differential
effects
on
the
GABAinduced
chloride
currents
in
human
embryonic
kidney
cells,
depending
upon
the
subunit
combinations
of
the
GABA
A­
receptor­
chloride
channel
complex.
The
current
molecular
biological
evidence
indicates
this
complex
to
normally
be
a
pentameric
protein
comprised
of
five
subunits
(",
$,
(,
*,
D)
in
various
combinations.
Using
the
whole
cell
variation
of
the
patch
clamp
technique,
the
EC
50
values
for
GABA
induction
of
chloride
current
were
estimated
as
9.8
:
M
for
the
"
1$
2
(
2s
combination,
2.0
:
M
for
the
"
1$
2
combination,
and
3.0
:
M
for
the
"
6$
2
(
2s
combination.
When
co­
applied
with
GABA,
dieldrin
(
1
to
3
:
M)
produced
mixed
effects:
initial
transient
enhancement,
followed
by
suppression,
was
observed
for
the
GABA­
induced
chloride
currents
in
the
"
1$
2
(
2s
and
"
6$
2
(
2s
combinations.
However,
suppression
alone
was
observed
in
the
"
1$
2
combination,
indicating
that
the
(
subunit
is
necessary
for
dieldrin
enhancing
effects.
The
EC
50
values
for
dieldrin's
suppression
of
GABA
induced
current
were
estimated
as
being
2.1
:
M
for
the
"
1$
2
(
2s
combination,
2.8
:
M
for
the
"
1$
2
combination,
and
1.0
:
M
for
the
"
6$
2
(
2s
combination.
The
authors
concluded
that
dieldrin­
induced
suppression
of
chloride
current
did
not
require
specific
subunit
combinations,
at
least
among
the
three
combinations
tested.

Using
the
radioligand,
t­
35S­
butyl­
bicyclophosphorothionate
(
35S­
TBPS),
Brannen
et
al.
(
1998)
demonstrated
that
dieldrin
could
interfere
with
GABA
receptor
binding.
The
authors
observed
that
injection
of
dieldrin
(
10
mg/
kg
bw/
day)
to
pregnant
dams
from
days
E12
through
E17
(
embryonic
days
12
to
17)
caused
a
significant
reduction
in
the
amount
of
(­
35S­
TBPS
binding
to
the
GABA
A
receptor
in
E17
brainstem
when
compared
to
vehicle­
injected
controls.
However,
dieldrin
treatment
caused
no
significant
effect
on
radioligand
binding
in
the
"
rest
of
the
brain."

Johns
et
al.
(
1998)
reported
findings
that
suggest
the
prenatal
binding
of
dieldrin
to
GABA
receptor
might
alter
the
expression
of
GABA
A
subunit
mRNAs
in
adult
offspring,
and
thus
might
explain
the
postnatal
behavioral
changes
observed
in
adult
offspring.
Pregnant
rats
were
injected
intraperitoneally
daily
during
E12
through
16
with
0,
5,
or
10
mg/
kg
bw
dieldrin.
On
postnatal
day
56,
testing
of
these
adult
offsprings
in
the
elevated
plus
maze
suggested
that
the
dieldrin­
treated
males
spent
less
time
in
the
closed
arms
of
the
elevated
maze,
indicating
a
lower
level
of
anxiety.
Consistent
with
these
behavioral
effects,
35STBPS­
binding
and
several
GABA
A
subunit
mRNA
levels
were
elevated
in
regions
of
the
brain
from
the
adult
offspring
of
dieldrintreated
dams.
7­
30
Aldrin/
Dieldrin
 
February
2003
In
an
in
vitro
study,
Liu
et
al.
(
1997)
studied
the
mechanisms
involved
in
the
neurotoxic
effect
of
dieldrin
on
(­
aminobutyric
acid
(
GABA
A)
receptors.
Using
embryonic
day
14
(
E14)
rat
brain
stem
cell
cultures,
the
authors
examined
the
effects
of
dieldrin
on
the
expression
of
five
GABA
receptor
mRNA
subunits
("
1,
$
3,
(
1,
(
2L,
(
2S).
Following
a
48­
hour
treatment
of
brain
stem
cells
with
10
:
M
dieldrin,
GABA
A
receptor
subunit
mRNA
levels
were
found
to
be
differentially
regulated
from
those
of
control
cultures.
Levels
of
$
3
subunits
were
significantly
increased
(
300%;
p<
0.05)
by
dieldrin,
whereas
expression
of
(
2S
and
(
2L
transcripts
were
decreased
by
50
and
40%,
respectively
(
p
<
0.05).
The
levels
of
"
1
and
(
1
subunits,
as
well
as
the
ratio
of
(
2S
to
(
2L,
were
not
significantly
affected
by
dieldrin
treatment.
As
the
evidence
in
general
suggested
a
correlation
between
gene
expression
and
receptor
function,
the
altered
expression
of
GABA
receptor
subunit
mRNAs
by
prenatal
dieldrin
exposure
may
affect
the
functional
properties
of
the
GABA
A
receptor
in
the
developing
brain.

Liu
et
al.
(
1998)
also
studied
the
effect
of
prenatal
in
vivo
exposure
to
dieldrin
on
the
expression
in
the
rat
fetal
brain
stem
of
the
five
GABA
A
receptor
subunit
mRNAs
("
1,
$
3,
(
1,
(
2L,
(
2S).
Pregnant
rats,
intraperitoneally
administered
dieldrin
at
1
mg/
kg
bw/
day
from
E12
to
E17,
evidenced
a
decrease
in
the
mRNA
levels
of
the
"
1,
$
3,
and
(
1
subunits,
but
not
of
those
for
(
2S
or
(
2L.
Again,
it
was
speculated
that
altered
expression
of
these
subunit
mRNAs
might
impact
the
functional
properties
of
the
GABA
A
receptor,
and
thus
GABA­
mediated
behaviors.

In
addition
to
its
effects
on
GABA­
ergic
neurons,
dieldrin
might
also
perturb
or
interact
with
dopaminergic
neurons.
In
an
in
vitro
study,
Sanchez­
Ramos
et
al.
(
1998)
examined
the
effects
of
a
24­
or
48­
hour
treatment
with
dieldrin
(
0.01
to
100
:
M)
on
primary
cultures
of
mesencephalic
neurons
isolated
from
the
fetal
brains
of
Sprague­
Dawley
rats
or
C
57/
BL
mice.
Toxicities
toward
dopaminergic
and
GABA­
ergic
neurons
were
assessed
by
determining
the
survival
of
tyrosine
hydroxylase­
immunoreactive
(
TH­
ir)
cells
and
glutamate
decarboxylase
(
GAD)­
ir
neurons,
respectively.
Dieldrin
exposure
for
24
hours
resulted
in
a
dose­
dependent
decrease
in
the
survival
of
TH­
ir
cells
from
rat
mesencephalic
cultures,
with
50%
relative
cell
survival
occurring
at
12
:
M.
The
24­
hour
dieldrin
treatment
also
produced
a
dose­
dependent
decrease
in
TH­
ir
cell
survival
in
mouse
mesencephalic
cultures,
with
75%
relative
cell
survival
occurring
at
10
:
M.
In
general,
the
toxic
effects
of
dieldrin
were
reported
to
be
more
severe
for
TH­
ir
neurons
than
for
GABA­
ergic
neurons.
Microscopic
changes
in
neurons
treated
with
dieldrin
were
observed
in
TH­
ir
cells,
such
as
diminished
numbers
and
lengths
of
neurites,
and
rounded
cell
bodies,
as
opposed
to
the
polygonal
or
spindle
form
found
in
control
neurons.
Consistent
with
the
cell
survival
effects,
dopamine
uptake
was
impaired
by
lower
concentration
of
dieldrin
than
was
GABA
uptake
(
the
EC
50
for
DA
uptake
was
7.98
:
M,
compared
to
43
:
M
for
GABA
uptake)
suggesting
that
dieldrin
had
a
greater
functional
effect
on
dopaminergic
neurons
than
on
GABA­
ergic
neurons.
Finally,
the
authors
concluded
that
the
greater
toxicity
of
dieldrin
on
dopaminergic
neurons
might
contribute
to
the
Parkinsonism
effects
observed
in
workers
exposed
to
pesticides,
such
as
dieldrin.

Chatterjee
et
al.
(
1992)
reported
a
number
of
estrogenic
effects
following
subcutaneous
administration
of
aldrin
(
1
mg/
kg
bw/
day)
for
3
days
to
groups
(
8/
group)
of
young
(
22­
day
old)
or
ovariectomized
adult
(
90­
day
old)
female
Wistar
rats.
Aldrin
exposure
caused
significant
increases
in
both
young
and
old
rats,
as
compared
to
their
respective
control
groups,
for
each
of
7­
31
Aldrin/
Dieldrin
 
February
2003
the
following
parameters
indicative
of
positive
estrogenic
effects:
uterine
weight,
endometrial
gland
thickness
and
proliferation,
and
the
staining
of
periodic
acid­
Schiff
positive
substance
in
the
uterus.
In
ovariectomized
adult
rats,
aldrin
exposure
also
induced
a
persistent
vaginal
estrus,
as
compared
to
the
constant
diestrus
seen
in
controls.

Several
mechanistic
studies
have
also
been
conducted
to
evaluate
the
estrogenic
activity
of
dieldrin.
Wade
et
al.
(
1997)
examined
the
possible
estrogenic
effects
of
dieldrin
in
vivo
by
measuring
the
estrogen
binding
activity,
peroxidase
activity,
and
uterine
weight
in
young
female
Spargue­
Dawley
rats,
and
in
vitro
by
measuring
the
cell
proliferation
activity
in
MCF­
7
cells
(
human
breast
cancer
cells).
Dieldrin
treatment
(
2
to
10
:
M)
could
competitively
inhibit
the
binding
of
3H­
17$­
estradiol
(
E
2)
to
estrogen
receptors
in
the
rat
uterus,
indicating
the
similarities
between
the
two
compounds.
The
authors
observed
that
the
intraperitoneal
administration
of
dieldrin
at
a
dose
of
3
mg/
kg
bw
to
young
female
rats
during
the
period
of
18
to
21
days
after
birth,
produced
no
changes
in
uterotrophic
activity
(
uterine
weight,
peroxidase
activity,
estrogen
receptor
number,
and
progesterone
receptor
number).
In
contrast
to
the
lack
of
these
specific
in
vivo
estrogenic
effects,
dieldrin
caused
a
positive
response
in
the
in
vitro
test;
treatment
of
cultured
MCF­
7
cells
with
50
:
M
dieldrin
resulted
in
a
3.4­
fold
increase
in
cell
proliferation,
as
compared
to
control
cells.
Wade
et
al.
(
1997)
also
reported
that
dieldrin
lacked
any
synergistic
effects
in
estrogenic
activity
when
tested
with
endosulfan
in
both
in
vitro
and
in
vivo
assays.
The
weak
in
vivo
estrogenic
response
of
dieldrin
is
not
likely
attributable
to
study
design
limitations,
as
the
positive
control,
diethylstilbesterol,
produced
significant
estrogenic
effects.

In
another
study,
Soto
et
al.
(
1994)
examined
the
estrogenic
effects
of
dieldrin
in
vitro
by
measuring
its
proliferative
effects
on
MCF­
7
cells.
Dieldrin
treatment
(
1.0
pM
to
10
:
M)
of
MCF­
7
cells
produced
a
significant
increase
in
the
proliferation
capacity
only
at
the
highest
concentration.
The
relative
proliferative
efficiency
for
dieldrin
at
10
:
M
was
54.89%
that
of
estradiol,
which
induced
its
maximum
level
of
proliferation
(
i.
e.,
100%
relative
proliferation)
at
a
concentration
of
only
10
pM
(
1
×
10­
6
of
the
tested
dieldrin
concentration).
This
indicates
the
relatively
very
weak
estrogenic
effect
of
dieldrin.

Ramamoorthy
et
al.
(
1997)
investigated
the
possible
estrogenic
activity
of
dieldrin
using
a
series
of
molecular
biology
assays:
estrogen
binding
activity
and
estrogen
effects
in
21­
day
old
B6C3F
1
mouse
uterus,
estrogen­
mediated
proliferation
in
MCF­
7
cells,
and
reporter
gene
assays
in
yeast
cells
transformed
with
mouse
or
human
estrogen
receptor
genes.
They
observed
that
dieldrin
did
not
bind
to
the
estrogen
receptor
in
mouse
uterus
or
MCF­
7
cells
in
a
competitive
manner;
produced
no
estrogen­
dependent
effects,
such
as
increase
in
uterine
wet
weight
or
progesterone
binding
in
uterus
excised
from
mice
(
treated
intraperitoneally
with
2.5
to
60
:
mol/
kg
bw/
day
for
3
days);
did
not
induce
MCF­
7
cell
proliferation
at
concentrations
ranging
from
10­
8
to
10­
5
M;
and
produced
minimal
induction
of
the
reporter
gene
activity
at
concentrations
of
up
to
2.5
×
10­
5
M
or
1
×
10­
4
M
in
yeast
cells
transformed
with
either
mouse
estrogen
receptor
or
human
estrogen
receptor,
respectively.
The
negative
findings
reported
in
this
study
regarding
dieldrin's
estrogenic
effects
are
not
attributable
to
study
design
limitations,
as
the
positive
controls
(
17$­
estradiol
and
diethylstilbesterol)
had
a
strong
estrogenic
effects
in
these
assays.
Overall,
the
authors
suggested
that
dieldrin
produced
only
minimal
estrogenic
7­
32
Aldrin/
Dieldrin
 
February
2003
effects
when
tested
alone,
and
moreover,
when
examined
with
toxaphene,
exhibited
no
synergistic
effects.

In
contrast
to
the
studies
of
Ramamoorthy
and
colleagues,
Arnold
et
al.
(
1996)
reported
that
dieldrin
might
have
estrogenic
effects
when
tested
alone,
or
possess
synergistic
effects
when
combined
with
endosulfan.
This
was
demonstrated
by
the
induced
expression
of
reporter
gene
activity
in
yeast
or
baculovirus
(
insect)
cells
that
had
been
transformed
with
human
estrogen
receptors.

Using
an
in
vitro
biomembrane
assay,
Demetrio
et
al.
(
1998)
investigated
the
effects
of
incubating
phosphatidylcholine
and
dimyristoylphosphatidylcholine
(
DMPC)
with
aldrin
concentrations
of
up
to
100
:
M
for
18
to
20
hours,
over
the
temperature
range
12
to
40
°
C.
They
reported
a
decrease
in
the
fluidity
of
the
lipid
bilayers,
as
measured
by
changes
in
fluorescence
polarization
of
DPH
(
1,6­
diphenyl­
1,3,5­
hexatriene)
and
of
its
propionic
acid
derivative
DPHPA
which
indicated
fluidity
changes
in
the
bilayer
core
and
in
the
outer
regions
of
the
bilayer,
respectively.
Although
these
membrane
fluidity
changes
may
alter
membrane
functions,
it
is
not
known
whether
they
occur
in
vivo.

Wright
et
al.
(
1972)
reported
that
within
1
week
of
exposing
rats
or
mice
to
8
or
1.6
mg/
kg
bw/
day
of
dieldrin,
respectively,
increases
were
observed
in
liver
cell
cytoplasmic
vacuoles,
smooth
endoplasmic
reticulum,
microsomal
protein,
and
mixed­
function
oxidase
activity.
They
also
reported
similar
effects
in
dogs
after
4
weeks
of
exposure
to
2
mg/
kg
bw/
day
of
dieldrin.
Exposure
of
monkeys
to
concentrations
as
high
as
0.1
mg/
kg
bw/
day
of
dieldrin
for
up
to
6
years
also
produced
increased
mixed­
function
oxidase
activity
and
cytochrome
P­
450
content
in
liver
cells
(
Wright
et
al.,
1972,
1978).

Finally,
in
vitro
studies
have
indicated
that
concentrations
of
aldrin
as
low
as
5.0
to
6.0
:
g/
ml
can
inhibit
gap­
junctional
intercellular
communication
among
human
teratocarcinoma
cells
(
Zhong­
Xiang
et
al.,
1986)
and
metabolic
cooperation
among
Chinese
hamster
cells
(
Kurata
et
al.,
1982).
Similarly,
dieldrin
has
also
been
reported
to
inhibit
intercellular
communication/
metabolic
cooperation
among
human
teratocarcinoma
cells
(
Wade
et
al.,
1986;
Zhong­
Xiang
et
al.,
1986),
Chinese
hamster
cells
(
Kurata
et
al.,
1982),
and
Syrian
hamster
embryo
cells
(
Mikalsen
and
Sanner,
1993).

7.3.5
Structure­
Activity
Relationship
Four
compounds
structurally
related
to
aldrin
and
dieldrin­
chlordane,
heptachlor,
heptachlor
epoxide,
and
chlorendic
acid­
have
induced
malignant
liver
tumors
in
mice;
chlorendic
acid
has
also
induced
liver
tumors
in
rats
(
USEPA,
1993a,
b).
7­
33
Aldrin/
Dieldrin
 
February
2003
7.4
Hazard
Characterization
7.4.1
Synthesis
and
Evaluation
of
Noncancer
Effects
Acute
exposures
to
aldrin
and/
or
the
metabolic
product,
dieldrin,
could
cause
neurotoxic
effects
in
humans
characterized
by
hyperirritability,
convulsions,
and
coma
(
Jager,
1970;
Spiotta,
1951;
ACGIH,
1984).
Cardiovascular
effects,
such
as
tachycardia
and
elevated
blood
pressure,
may
occur
subsequent
to
convulsions
(
Black,
1974).

Evidence
suggesting
that
children
would
experience
the
neurotoxic
toxic
effects
upon
acute
exposure
to
aldrin
and/
or
dieldrin
is
limited.
In
many
instances,
children
may
be
more
sensitive
than
adults
as
a
result
of
their
developing
organ
systems
(
e.
g.,
nervous
system)
and
metabolic
detoxication
capacities
(
Hayes,
1982;
ATSDR,
2000).
Long­
term
effects
of
aldrin/
dieldrin
in
children
have
not
been
studied.

Occupational
studies
suggest
that
workers
involved
in
the
manufacture
or
application
of
aldrin/
dieldrin
have
increased
dieldrin
levels
in
plasma
(
up
to
250
ng/
mL).
From
the
plasma,
the
aldrin
and
dieldrin
could
be
distributed
and
stored
in
adipose
tissue
(
Nair
et
al.,
1992).

Aldrin
and
dieldrin
are
quite
toxic,
as
they
have
low
acute
toxicity
values
when
tested
in
animals
(
LD
50
values
of
generally
#
100
mg/
kg)
(
Borgmann
et
al.,
1952;
Gaines,
1960;
Treon
et
al.,
1952;
Lu
et
al.,
1965).
Common
acute
or
subchronic
neurotoxic
effects
observed
in
animals
are
characterized
by
increased
irritability,
salivation,
hyperexcitability,
tremors
followed
by
convulsions,
loss
of
body
weight,
depression,
prostration,
and
death
(
Borgmann
et
al.,
1952;
Walker
et
al.,
1969;
Wagner
and
Greene,
1978;
Wooley
et
al.,
1985;
NCI,
1978;
Casteel,
1993).
These
symptoms
are
similar
to
those
observed
in
humans
exposed
to
aldrin
or
dieldrin
(
Jager,
1970;
Spiotta,
1951;
ACGIH,
1984;
ATSDR,
2000).
In
addition,
cardiovascular
effects,
such
as
tachycardia
and
elevated
blood
pressure,
may
occur
in
humans
subsequent
to
convulsions
(
Black,
1974).

Evidence
suggests
that
short­
term
or
subchronic
oral
exposure
to
aldrin
at
dietary
concentrations
of
300
ppm
in
rats
and
80
ppm
in
mice
could
result
in
high
mortality
rates
(
Treon
and
Cleveland,
1955;
NCI,
1978).

Subchronic
exposure
of
B6C3F
1
mice
to
dieldrin
caused
no
significant
effects
in
body
weight
gains,
food
consumption,
or
water
consumption,
but
could
increase
relative
liver
weights
(
liver
weight
to
body
weight
ratios),
and
promote
hepatic
lesions
induced
by
the
hepatic
carcinogen
diethylnitrosamine.
These
dieldrin­
induced
hepatic
effects
may
be
specific
to
mice,
as
they
were
not
observed
in
rats
(
Kolaja
et
al.,
1996a,
b).

Aldrin/
dieldrin
exposure
has
been
shown
to
produce
developmental
and
reproductive
toxic
effects.
In
a
3­
generation
reproduction
study
conducted
in
rats,
a
reduction
in
the
pregnancy
rate
was
reported
(
Treon
and
Cleveland,
1955).
Prenatal
exposure
to
aldrin
also
appears
to
have
caused
a
reduction
in
pup
survival
in
dogs
(
Deichmann
et
al.,
1971).
An
increase
in
fetal
deaths,
a
decrease
in
live
fetal
weight,
and
increased
incidences
of
webbed
foot,
cleft
7­
34
Aldrin/
Dieldrin
 
February
2003
palate,
and
open
eye
were
reported
in
Syrian
golden
hamsters
and
CD1
mice
that
were
exposed
to
dieldrin
prenatally
(
Ottolenghi
et
al.,
1974).
Prenatal
exposure
to
dieldrin
may
also
cause
maternal
toxic
effects,
such
as
the
increase
in
maternal
mortality
reported
in
mice
(
Virgo
and
Bellward
1975).

However,
evidence
from
several
animal
studies
does
not
indicate
that
reproductive
effects
such
as
changes
in
fertility,
fecundity,
length
of
gestation,
or
perinatal
mortality
would
be
likely
to
result
from
exposure
at
environmental
levels
to
either
aldrin
or
dieldrin
(
Kitselman,
1953;
Good
and
Ware,
1969;
Harr
et
al.,
1970;
Coulston
et
al.,
1980).

Experimental
evidence
suggesting
that
dieldrin
may
be
capable
of
producing
estrogenic
effects
is
not
consistent
(
Ramamoorthy
et
al.,
1997;
Arnold
et
al.,
1996;
Wade
et
al.,
1997).
Chatterjee
et
al.
(
1992)
found
that
administration
of
dieldrin
to
rats
increased
uterine
weight,
endometrial
gland
thickness
and
proliferation,
and
the
level
of
staining
for
periodic
acid­
Schiff
positive
substances
in
the
uterus
in
a
fashion
similar
to
that
seen
for
estrogen.
Soto
et
al.
(
1994)
reported
weak
estrogenic­
like
effects
for
dieldrin
with
respect
to
its
enhancement
of
the
proliferation
of
human
breast
cancer
cells.

Chronic
feeding
of
aldrin
(
1
to
10
mg/
kg
bw/
day)
to
animals
has
in
general
produced
high
mortality
effects
(
Treon
and
Cleveland,
1955;
Fitzhugh
et
al.,
1964;
NCI,
1978;
Kitselman
and
Borgmann,
1952).
Similar
results
were
also
reported
for
dieldrin
exposure
in
mice
(
Walker
et
al.,
1972).

Increased
liver
weights
and
liver­
to­
body
weight
ratios
were
observed
consistently
in
rats
chronically
exposed
to
aldrin
(
Treon
and
Cleveland,
1955;
Deichmann
et
al.,
1970;
Fitzhugh
et
al.,
1964),
and
in
rats,
dogs,
and
mice
exposed
to
dieldrin
(
Fitzhugh
et
al.,
1964;
Walker
et
al.,
1969;
Walker
et
al.,
1972).

Neurotoxic
effects
similar
to
those
seen
after
acute
exposure,
such
as
tremors
and
convulsions,
have
also
been
reported
in
long­
term
oral
studies
of
animals
exposed
to
aldrin
(
Fitzhugh
et
al.,
1964;
Deichmann
et
al.,
1970)
or
dieldrin
(
Fitzhugh
et
al.,
1964).

Chronic
oral
exposure
of
rats
to
aldrin
has
also
caused
some
histopathological
alterations
in
the
liver,
which
were
characterized
by
enlarged
centrilobular
hepatic
cells
having
increased
cytoplasmic
oxyphilia
and
peripheral
migration
of
basophilic
granules
(
Fitzhugh
et
al.,
1964).

No
studies
were
obtained
that
examined
the
toxic
effects
of
aldrin
or
dieldrin
in
animals
following
inhalation
exposure,
and
only
very
limited
information
was
found
regarding
dermal
exposure.

7.4.2
Synthesis
and
Evaluation
of
Carcinogenic
Effects
Long­
term
follow­
up
studies
suggest
that
standardized
mortality
rates
for
all
causes
of
death
in
workers
employed
in
pesticide
manufacturing
plants
are
significantly
lower
than
the
7­
35
Aldrin/
Dieldrin
 
February
2003
corresponding
national
mortality
rates
(
de
Jong,
1991;
de
Jong
et
al.,
1997;
Ditraglia
et
al.,
1981;
Brown,
1992;
Amaoteng­
Adjepong
et
al.,
1995).
Slight
increases
in
the
incidence
of
rectal
and
liver
cancers
have
been
observed
in
the
aldrin/
dieldrin
exposed
groups,
but
they
are
not
robust
or
dose­
dependent
(
de
Jong
et
al.,
1997;
Ditraglia
et
al.,
1981).
Most
of
the
results
from
the
various
occupational
studies
on
the
human
health
effects
of
aldrin/
dieldrin
exposure
are
complicated
to
some
degree
by
the
simultaneous
exposure
to
other
pesticides;
most
plants
were
also
involved
in
the
manufacture
of
other
pesticides,
and
the
association
of
adverse
human
health
effects
with
aldrin/
dieldrin
contact
is
weakened
by
the
lack
of
adequate
exposure
assessment.

Several
lines
of
evidence
suggest
that
chronic
exposure
to
aldrin/
dieldrin
selectively
increases
the
incidence
of
liver
cancer
in
several
different
strains
of
mice
(
Meierhenry
et
al.,
1983;
Davis
and
Fitzhugh,
1962,
Davis,
1965;
Epstein,
1975a,
b;
NCI,
1978;
Thorpe
and
Walker,
1973;
Walker
et
al.,
1972;
Tennekes
et
al.,
1981).
These
results
were
not
observed
in
several
strains
of
rats
that
have
also
been
tested
(
Treon
and
Cleveland,
1955;
Fitzhugh
et
al.,
1964;
Song
and
Harville,
1964;
Walker
et
al.,
1969;
Deichmann
et
al.,
1970;
NCI,
1978).
Evidence
from
a
single
rat
study
(
NCI,
1978)
suggesting
possible
increases
in
the
incidences
of
follicular
cell
adenoma
and
carcinoma
of
the
thyroid
and
of
cortical
adenoma
of
the
adrenal
gland
after
chronic
aldrin
exposure
has
not
been
supported
by
other
studies.
It
must
be
kept
in
mind,
however,
that
a
number
of
these
studies
has
been
deemed
inadequate
tests
for
carcinogenicity
due
to
a
variety
of
significant
study
limitations.

Seven
studies
that
collectively
utilized
4
strains
of
rats,
which
were
fed
0.1
to
285
ppm
dieldrin
for
durations
varying
from
80
weeks
to
31
months,
did
not
produce
positive
results
for
carcinogenicity
(
Treon
and
Cleveland,
1955;
Fitzhugh
et
al.,
1964;
Song
and
Harville,
1964;
Walker
et
al.,
1969;
Deichmann
et
al.,
1970;
NCI,
1978).
Three
of
these
studies
used
Osborne­
Mendel
rats,
two
studies
used
Carworth
rats,
and
one
each
used
Fischer
344
and
Holtzman
strains.
As
noted
above
for
aldrin,
only
three
of
the
seven
dieldrin
studies
were
considered
adequate
in
design
and
conduct
(
USEPA,
1987,
1993b).
The
others
used
too
few
animals,
had
unacceptably
high
levels
of
mortality,
were
too
short
in
duration,
and/
or
had
inadequate
pathology
examination
or
reporting.

The
status
of
aldrin
and
dieldrin
as
genotoxins
is
somewhat
equivocal.
Summarizing
the
studies
reviewed
in
this
document
by
certain
genotoxicity
endpoint
categories,
the
assays
performed
on
one
or
both
of
these
chemicals
have
produced
the
following
responses:

bacterial
gene
mutation:
21
(!)
1
(+)
fungal
gene
mutation/
conversion:
4
(!)
1
(+)
in
vitro
mammalian
cell
gene
mutation:
1
(+)
mammalian
host­
mediated
bacterial
gene
mutation:
1
(!)
in
vivo
gene
mutation
­
insects:
2
(!)
in
vitro
chromosome
damage/
aneuploidy:
2
(!)
3
(+)
in
vivo
chromosome
damage/
aneuploidy:
6
(!)
4
(+),
3
(?+)
in
vitro
SCE:
1
(+)
in
vivo
SCE:
1
(!)
1
(?+)
bacterial/
plasmid
DNA
damage:
2
(!)
in
vitro
mammalian
cell
DNA
damage:
4
(!),
2
(?!)
3
(+)
in
vitro
cell
transformation:
1
(!)
7­
36
Aldrin/
Dieldrin
 
February
2003
While
the
preponderance
of
these
assay
results
are
negative,
some
of
the
in
vitro
assays
failed
to
employ
some
form
of
exogenous
metabolic
activation,
such
as
S9
mix.
Based
on
these
data,
it
is
currently
difficult
to
reject
at
least
the
possibility
that
aldrin/
dieldrin
can
interact
with
chromosomes
or
induce
DNA
damage.
However,
as
suggested
in
the
following
section,
some
or
all
of
aldrin/
dieldrin's
apparent
genotoxicity
may
indirect
or
reflect
epigenetic
mechanisms.

7.4.3
Mode
of
Action
and
Implications
in
Cancer
Assessment
There
have
been
several
mechanistic
studies
conducted
to
explain
the
selective
hepatocarcinogenic
effects
of
aldrin
and
dieldrin
in
mice.
Several
studies
suggest
that
these
chemicals
can
induce
at
least
hepatic
tumors
in
mice,
but
are
much
less
likely
to
do
so,
if
at
all,
in
rats.
Although
the
mechanisms
responsible
for
this
species
specificity
are
not
fully
understood,
accumulating
evidence
indicates
that
increased
hepatic
DNA
synthesis
and
oxidative
stress
may
be
involved.

Kolaja
et
al.
(
1996a)
observed
an
increased
DNA
labeling
index
in
the
centrilobular
region
of
liver
in
male
B6C3F
1
mice,
but
not
in
male
Fisher
344
rats,
as
early
as
7
or
14
days
after
exposure
to
dieldrin
at
concentrations
of
3.0
or
10.0
mg/
kg
diet.
Increases
in
the
liver
DNA
labeling
index
in
mice,
but
not
in
rats,
were
also
reported
by
Kolaja
et
al.
(
1995)
for
the
highdose
groups
when
animals
(
5/
species/
dose)
were
fed
with
dieldrin
in
the
diet
at
0
(
control),
0.1,
1.0,
or
10.0
mg
dieldrin/
kg
diet
for
7
or
14
days.
In
a
subsequent
study,
Kolaja
et
al.
(
1996b)
described
a
selective
promotion
of
hepatic
focal
lesions
and
an
increase
in
DNA
labeling
at
the
highest
dose
in
male
B6C3F
1
mice,
but
not
in
male
Fisher
344
rats.
Groups
of
animals
(
5
animals/
species/
dose)
were
treated
with
the
hepatic
carcinogen,
diethylnitrosamine
(
150
mg/
kg
bw/
week,
2x
for
rats;
25
mg/
kg
bw/
week,
8x
for
mice),
prior
to
the
administration
of
dieldrin
at
0.1,
1.0,
or
10.0
mg/
kg
diet
for
7
days.

In
vivo
experiments
by
Bachowski
et
al.
(
1997)
demonstrated
the
following
in
B6C3F
1
mice,
but
not
in
F344
rats,
upon
feeding
10.0
mg
dieldrin/
kg
diet
for
up
to
540
days:
1)
increased
production
of
2,3­
DHBA
(
2,3­
dihydroxybenzoic
acid,
a
marker
used
for
measuring
oxidative
stress)
in
hepatocytes
and
their
microsomes;
2)
elevated
production
of
MDA
(
malondialdehyde,
a
marker
for
oxidative
damage
to
lipids)
in
liver
and
urine;
3)
increased
OH8dG
(
8­
hydroxy­
2'­
deoxyguanosine,
a
marker
for
oxidative
damage
to
DNA)
levels
in
urine;
and
4)
decreased
hepatic
vitamin
E
("­
tocopherol)
content.
The
authors
concluded
that
oxidative
stress
mechanisms
may
be
involved
in
the
mediation
of
dieldrin­
induced
hepatic
DNA
synthesis
that
is
observed
in
mice,
but
not
in
rats.

In
a
more
recent
study,
Bachowski
et
al.
(
1998)
also
examined
the
in
vivo
association
between
dieldrin­
induced
hepatic
DNA
synthesis
and
oxidative
damage
to
lipids
(
MDA),
DNA
(
OH8dG),
or
levels
of
nonenzymatic
antioxidants
(
ascorbic
acid,
glutathione,
vitamin
E)
in
male
B6C3F
1
mice
and
F344
rats
that
were
fed
dieldrin
(
0.1,
1.0,
or
10
mg/
kg
diet)
for
up
to
90
days.
Consistent
with
the
increase
in
hepatic
DNA
synthesis
induced
by
dieldrin
treatment,
decreases
in
hepatic
and
serum
vitamin
E
levels
("­
tocopherol),
and
increases
in
hepatic
MDA
and
urinary
MDA
and
OH8dG
levels,
were
observed
in
mice.
In
contrast,
these
effects
were
less
dramatic
or
7­
37
Aldrin/
Dieldrin
 
February
2003
not
observed
in
rats,
which
may
have
been
protected
by
higher
basal
levels
of
vitamin
E
(
and
vitamin
C)
in
their
hepatic
tissue.

Stevenson
et
al.
(
1995)
found
indirect
evidence
for
an
oxidative
stress
mechanism
when
they
measured
a
partial
protective
effect
of
vitamin
E
in
ameliorating
the
dieldrin­
induced
hepatic
DNA
synthesis
observed
in
B6C3F
1
mice
(
fed
dieldrin
at
10
mg/
kg
diet
for
28
days).
However,
supplementation
of
the
diet
with
vitamin
C
(
another
antioxidant)
at
up
to
400
mg/
kg
diet
resulted
in
only
an
inconsistent
reduction
in
dieldrin­
induced
hepatic
DNA
labeling.

Kolaja
et
al.
(
1998)
also
reported
that
supplementation
of
the
diet
with
vitamin
E
(
450
mg/
kg
diet)
for
up
to
60
days
blocked
the
dieldrin
(
10
mg/
kg
diet)
treatment
effects
of
increased
hepatic
focal
lesion
volume,
focal
lesion
number,
and
focal
lesion
DNA
labeling
index
that
were
observed
in
mice
pre­
treated
with
the
hepatic
carcinogen,
diethylnitrosamine.

Bauer­
Hofmann
et
al.
(
1992)
analyzed
the
frequency
and
pattern
of
c­
Ha­
ras
protooncogene
mutations
at
codon
61
in
polymerase
chain
reaction­
amplified
DNA
taken
from
glucose­
6­
phosphatase
deficient
(
G6P!)
hepatic
lesions
in
groups
of
male
C3H/
He
mice;
groups
had
received
either
10
ppm
dieldrin,
500
ppm
phenobarbital
(
PB),
or
no
treatment
in
the
diet
for
52
weeks.
The
incidence
of
G6P!
hepatic
lesions
was
reported
to
increase
from
41%
(
15/
37)
in
the
control
group
to
67%
(
10/
15)
and
63%
(
10/
16)
in
the
dieldrin
and
PB
groups,
respectively.
The
corresponding
average
numbers
of
focal
lesions/
mouse
were
0.57,
1.5,
and
1.0.
Upon
DNA
analysis,
c­
Ha­
ras
mutations
were
observed
in
57%
(
12/
21)
of
the
lesions
from
the
control
group,
but
in
only
22%
(
5/
23)
and
25%
(
4/
16)
of
those
from
the
dieldrin
and
PB
groups,
respectively.
As
dieldrin
(
like
PB)
increased
the
frequency
of
c­
Ha­
ras
wild­
type,
but
not
mutated,
focal
hepatic
lesions,
the
authors
concluded
that
the
c­
Ha­
ras
mutations
had
likely
occurred
spontaneously
rather
than
as
a
result
of
dieldrin
treatment.
Further,
no
significant
differences
in
the
mutation
spectra
were
noticed
between
control
and
dieldrin
treated
mice,
the
most
prominent
class
of
mutation
being
C6A
transversion.
These
data
suggest
that
the
principal
role
for
dieldrin
in
liver
tumor
formation
may
be
one
of
promotion,
rather
than
initiation.

Based
on
existing
studies,
Stevenson
et
al.
(
1999)
have
also
suggested
that
aldrin/
dieldrin
exposure
induces
hepatocarcinogenesis
in
mice
through
non­
genotoxic
mechanisms
such
as
increased
production
of
reactive
oxygen
species
(
ROS)
in
mouse
hepatocytes
(
possibly
by
futile
cycling
of
P450
enzymes),
increased
hepatic
DNA
synthesis,
and
augmentation
of
tumorpromotional
effects,
rather
than
by
causing
point
mutations
or
otherwise
directly
interacting
with
DNA.
A
possible
mode
of
action
for
aldrin/
dieldrin
in
animals
is
depicted
in
Figure
7­
l.
Although
the
figure
depicts
aldrin/
dieldrin
induction
of
hepatic
DNA
synthesis
through
modulation
of
proto­
oncogene
expression
(
via
transcription
factors
such
as
Nf­
kB,
AP­
1,
etc.),
data
directly
relating
the
effects
of
aldrin/
dieldrin
exposure
to
protooncogene
expression
remain
to
be
established.

In
addition
to
mechanisms
that
involve
oxidative
stress
and
the
direct
promotion
of
cellular
proliferation,
the
previously
discussed
capacity
of
aldrin
and
dieldrin
to
inhibit
various
forms
of
in
vitro
intercellular
communication
in
both
human
and
animal
cells
may
be
significant
7­
38
Aldrin/
Dieldrin
 
February
2003
Neoplasia
Aldrin/
Dieldrin
ROS
Production
Modulation
of
Gene
Expression
(
NF­
kB,
AP­
1,
c­
Ha­
ras,
second
messengers)

Increased
S­
phase
DNA
synthesis
in
hepatocytes
Nongenotoxic
(
Yes)
Genotoxic
(
No)

Genetic
Instability
Increase
in
focal
lesion
size
Increased
cell
proliferation
of
spontaneously
initiated
hepatocytes
Mitosis
with
respect
to
their
in
vivo
effects
on
tumor
production
(
Kurata
et
al.,
1982;
Wade
et
al.,
1986;
Zhong­
Xiang
et
al.,
1986;
Mikalsen
and
Sanner,
1993).

7.4.4
Weight
of
Evidence
Evaluation
for
Carcinogenicity
Using
current
EPA
(
1986)
cancer
guidelines,
aldrin
and
dieldrin
are
classified
as
B2
carcinogens,
i.
e.
probable
human
carcinogens
with
little
or
no
evidence
of
carcinogenicity
in
humans
and
sufficient
evidence
in
animals
(
different
strains
of
mice).
With
inadequate
data
on
carcinogenic
effects
in
humans,
under
the
USEPA's
cancer
risk
assessment
guidelines
(
USEPA,
1996/
1999),
the
weight
of
evidence
indicates
that
aldrin
and
dieldrin
could
be
classified
as
rodent
carcinogens
that
are
"
likely
to
be
carcinogenic
to
humans
by
the
oral
route
of
exposure,
but
whose
carcinogenic
potential
by
the
inhalation
and
dermal
routes
of
exposure
cannot
be
determined
because
there
are
inadequate
data
to
perform
an
assessment."
This
characterization
is
based
on
the
tumor
effects
of
aldrin
and
dieldrin
observed
in
several
strains
of
mice
subsequent
to
oral
exposures
and
must
be
tempered
by
the
lack
of
evidence
for
significant
human
carcinogenicity
from
epidemiological
studies.
It
should
be
noted
that
the
USEPA
has
quantified
the
estimated
carcinogenic
risks
from
inhalation
exposure
to
aldrin
and
dieldrin
by
extrapolating
from
available
oral
exposure
route
data
(
USEPA,
1993a,
b).
Mechanistic
studies
performed
in
7­
39
Aldrin/
Dieldrin
 
February
2003
Figure
7­
1.
The
Possible
Mode
of
Action
of
Aldrin/
Dieldrin
on
Hepatocarcinogenesis
Adapted
from
Stevenson
et
al.
(
1999)

vitro
and
in
vivo
suggest
that
one
or
more
non­
genotoxic
modes
of
action
may
underlie
or
contribute
to
the
carcinogenic
potential
of
aldrin
and
dieldrin,
but
these
effects
are
not
completely
established,
or
can
a
role
for
genotoxic
mechanisms
confidently
be
eliminated
based
on
the
available
data
.
In
the
absence
of
adequate
data
to
fully
support
a
non­
linear
mechanism(
s)
of
tumor
formation,
the
quantitative
cancer
risk
assessment
of
aldrin
and
dieldrin
should
conservatively
be
conducted
using
the
linear­
default
model.

7.4.5
Sensitive
Populations
No
human
studies
were
obtained
that
adequately
address
the
effect
of
aldrin
and
dieldrin
on
sensitive
populations,
such
as
children.
Several
mechanistic
studies,
which
describe
the
prenatal
effects
of
aldrin/
dieldrin
on
GABA
receptor
malfunctions
and
on
subsequent
behavioral
impairment,
may
suggest
that
children
could
be
more
sensitive
to
aldrin
and
dieldrin
exposures
than
the
general
adult
population
(
Brannen
et
al.,
1998;
Liu
et
al.,
1998;
Johns
et
al.,
1998;
Castro
et
al.,
1992).
7­
40
Aldrin/
Dieldrin
 
February
2003
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epa.
gov/
ngispgm3/
iris/
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Aldrin
drinking
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advisory.
Washington,
DC:
USEPA,
Office
of
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Dieldrin
 
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2003
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S.
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USEPA,
Office
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USEPA.
1987.
U.
S.
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006.
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USEPA,
Office
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Environmental
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USEPA.
1986.
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with
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1988.
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1
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term
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Virgo,
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D.
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1977.
Effects
of
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dieldrin
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Virgo,
B.
B.
and
G.
D.
Bellward.
1975.
Effects
of
dietary
dieldrin
on
reproduction
in
the
Swiss­
Vancouver
(
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Biochem..
5:
440­
450
(
as
cited
in
ATSDR,
2000;
IPCS,
1989).

Wade,
M.
G.,
D.
Desaulniers,
K.
Leingartner,
and
W.
G.
Foster.
1997.
Interactions
between
endosulfan
and
dieldrin
on
estrogen­
mediated
processes
in
vitro
and
in
vivo.
Reprod.
Toxicol.
11:
791­
798.

Wade,
M.
H.,
J.
E.
Trosko,
and
M.
Schindler.
1986.
A
flourescence
photobleaching
assay
of
gap
junction­
mediated
communication
between
human
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Science
232:
525­
528
(
as
cited
in
GAP2000,
2000b).

Wade,
M.
J.,
J.
W.
Moyer,
and
C.
H.
Hine.
1979.
Mutagenic
action
of
a
series
of
epoxides.
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Res.
66(
4):
367­
371
(
as
cited
in
USEPA,
1993b;
GAP2000,
2000b).

Wagner,
S.
R.
and
F.
E.
Greene.
1978.
Dieldrin­
induced
alterations
in
biogenic
amine
content
of
rat
brain.
Toxicol.
Appl.
Pharmacol.
43:
45­
55
(
as
cited
in
ATSDR,
2000).

Walker,
A.
I.
T.,
E.
Thorpe,
and
D.
E.
Stevenson.
1972.
The
toxicology
of
dieldrin
(
HEOD).
I.
Long­
term
oral
toxicity
studies
in
mice.
Food
Cosmet.
Toxicol.
11:
415­
432
(
as
cited
in
USEPA,
1987,
1988,
and
1993b).
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53
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Dieldrin
 
February
2003
Walker,
A.
I.
T.,
D.
E.
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J.
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E.
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and
M.
Roberts.
1969.
The
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and
pharmacodynamics
of
dieldrin
(
HEOD):
Two­
year
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exposures
of
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Toxicol.
Appl.
Pharmacol.
15:
345­
373
(
as
cited
in
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1987,
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1993b).

Warnick
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L.
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1972.
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Environ.
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25:
265­
270
(
as
cited
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Witherup,
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J.
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1961.
Prolonged
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The
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Cincinnati,
OH
(
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1985.
Effects
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C.
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1978.
The
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of
prolonged
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of
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on
the
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R.
D.
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1972.
The
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of
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on
the
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of
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cells.
Food
Cosmet.
Toxicol.
10:
311­
332
(
as
cited
in
ATSDR,
2000).

Zhong­
Xiang,
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T.
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J.
E.
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C.
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1986.
Inhibition
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Toxicol.
Appl.
Pharmacol.
83:
10­
19
(
as
cited
in
GAP2000,
2000a,
b).
8­
1
Aldrin/
Dieldrin
 
February
2003
8.0
DOSE­
RESPONSE
ASSESSMENTS
8.1
Dose­
Response
for
Non­
Cancer
Effects
8.1.1
Reference
Dose
Determination
The
oral
Reference
Dose
(
RfD),
formerly
termed
the
Acceptable
Daily
Intake
(
ADI),
is
based
on
the
assumption
that
thresholds
exist
for
most,
if
not
all,
noncancer
toxic
effects.
In
general,
the
RfD
is
an
estimate
(
with
uncertainty
spanning
perhaps
an
order
of
magnitude)
of
a
daily
exposure
to
the
human
population
(
including
sensitive
subgroups)
that
is
likely
to
be
without
an
appreciable
risk
of
deleterious
effects
during
a
lifetime.
The
RfD
is
expressed
in
units
of
mg/
kg
bw/
day,
and
has
traditionally
been
derived
from
the
NOAEL
(
or
LOAEL)
identified
from
the
data
in
a
chronic
(
or
subchronic)
study,
divided
by
an
uncertainty
factor
composed
of
one
or
more
elements
defined
by
EPA
or
NAS/
OW
guidelines.

Aldrin
Choice
of
Principal
Study
and
Critical
Effect
The
rat
study
by
Fitzhugh
et
al.
(
1964),
designed
as
a
carcinogenesis
bioassay,
has
been
selected
to
serve
as
the
basis
for
the
Reference
Dose
principally
because
it
displayed
strength
in
histopathologic
analysis,
it
examined
a
wider
dose
range
(
0.5
to
150
ppm
in
the
diet)
when
compared
with
other
available
chronic
studies,
and
in
the
absence
of
a
reliable
NOAEL,
its
data
established
the
lowest
available
LOAEL.
The
database
is
fairly
extensive
and,
generally,
supportive
of
the
principal
study's
findings,
but
is
rated
medium
because
of
the
lack
of
NOELs.
Other
chronic
studies
in
rats
(
using
dietary
exposures
of
2.5
to
60
ppm)
and
dogs
have
also
demonstrated
aldrin's
toxic
effects
on
the
liver
(
Deichmann
et
al.,
1970;
Treon
and
Cleveland,
1955;
NCI,
1978;
the
dog
study
in
Fitzhugh
et
al.,
1964).

In
the
principal
study,
groups
of
24
rats
(
12/
sex)
were
fed
aldrin
in
the
diet
at
levels
of
0,
0.5,
2,
10,
50,
100,
or
150
ppm
for
2
years.
Liver
lesions
characteristic
of
chlorinated
insecticide
poisoning
were
observed
at
all
exposure
levels
of
aldrin.
These
lesions
were
characterized
by
enlarged
centrilobular
hepatic
cells,
with
increased
cytoplasmic
oxyphilia
and
peripheral
migration
of
basophilic
granules.
In
addition,
a
statistically
significant
increase
in
liver­
to­
body
weight
ratio
was
observed
at
all
dose
levels.
Kidney
lesions
at
the
highest
dose
levels
were
also
reported
and
survival
was
markedly
decreased
at
dose
levels
of
50
ppm
and
greater.
The
effect
and
no­
effect
levels
for
liver
toxicity
are
similar
to
those
reported
in
the
same
study
for
dogs
exposed
to
aldrin
in
the
diet
for
15
months
(
Fitzhugh
et
al.,
1964).
While
not
permitting
the
determination
of
a
NOAEL,
the
study
does
establish
a
LOAEL
at
the
lowest
aldrin
concentration
tested,
0.5
ppm.

RfD
Derivation
The
RfD
for
aldrin
was
derived
from
the
critical
effect
(
liver
toxicity)
that
it
induced
in
rats
during
a
2­
year
chronic
feeding
study
(
Fitzhugh
et
al.,
1964).
This
principal
study
reported
8­
2
Aldrin/
Dieldrin
 
February
2003
that
various
toxic
effects
occurred
in
the
liver
at
all
aldrin
concentrations
tested
(
0.5
to
150
ppm).
The
resulting
LOAEL
dietary
concentration
of
0.5
ppm
can
be
converted
to
a
dose
of
0.025
mg/
kg
bw/
day
by
using
an
equivalency
factor
of
1
ppm
in
the
diet
=
0.05
mg/
kg
bw/
day
based
on
the
food
consumption
rate
in
rats
as
described
by
Lehman
(
1959).

The
RfD
for
aldrin
is
calculated
as
follows:

RfD
=
=
0.000025
mg/
kg
bw/
day
(
rounded
to
3E­
5
mg/
kg
bw/
day)
0.025
mg
/
kg
bw
/
day
1000
where:

0.025
mg/
kg
bw/
day
=
LOAEL,
based
on
liver
toxicity
in
rats
exposed
to
aldrin
in
the
diet
for
2
years
1000
=
uncertainty
factor;
this
composite
uncertainty
factor
was
chosen
in
accordance
with
EPA
or
NAS/
OW
guidelines
in
which
uncertainty
factors
of
10
each
were
applied
to
extrapolate
from
rats
to
humans,
to
account
for
uncertainty
in
the
range
of
human
sensitivity
(
i.
e.,
to
protect
sensitive
human
subpopulations),
and
to
account
for
additional
uncertainty
because
the
study
identified
a
LOAEL
(
but
not
a
NOAEL).

Dieldrin
Choice
of
Principal
Study
and
Critical
Effect
The
study
by
Walker
et
al.
(
1969),
also
designed
as
a
carcinogenesis
bioassay,
has
been
selected
to
serve
as
the
basis
for
the
Reference
Dose
principally
because
it
was
fairly
extensively
reported,
the
exposure
period
was
of
chronic
duration,
NOAELs
were
determined,
and
it
is
generally
supported
by
other
toxicity
studies
of
dieldrin.

Walker
et
al.
(
1969)
administered
dieldrin
(
recrystallized,
99%
active
ingredient)
to
Carworth
Farm
"
E"
rats
(
25/
sex/
dose;
controls
45/
sex)
for
2
years
at
dietary
concentrations
of
0,
0.1,
1.0,
or
10.0
ppm.
Based
on
intake
assumptions
presented
by
the
authors,
these
dietary
levels
are
approximately
equivalent
to
0,
0.005,
0.05,
and
0.5
mg
dieldrin/
kg
bw/
day,
respectively.
Body
weight,
food
intake,
and
general
health
remained
unaffected
throughout
the
2­
year
period,
although
at
10.0
ppm
(
0.5
mg/
kg
bw/
day)
all
animals
became
irritable
and
exhibited
tremors
and
occasional
convulsions.
No
effects
were
observed
for
various
hematological
and
clinical
chemistry
parameters.
At
the
end
of
2
years,
females
fed
1.0
and
10.0
ppm
(
0.05
and
0.5
mg/
kg
bw/
day,
respectively)
had
increased
liver
weights
and
liver­
to­
body
weight
ratios
(
p
<
0.05).
Histopathological
examinations
revealed
liver
parenchymal
cell
changes,
including
focal
proliferation
and
focal
hyperplasia.
These
hepatic
lesions
were
considered
by
the
authors
to
be
characteristic
of
exposure
to
an
organochlorine
insecticide.
Based
on
these
toxic
effects
8­
3
Aldrin/
Dieldrin
 
February
2003
observed
in
the
liver,
the
LOAEL
was
identified
as
1.0
ppm
(
0.005
mg/
kg
bw/
day)
and
the
NOAEL
as
0.1
ppm
(
0.005
mg/
kg
bw/
day).

RfD
Derivation
The
RfD
for
dieldrin
was
derived
from
the
critical
effect
(
altered
absolute
and
relative
liver
weights
that
were
accompanied
by
histopathological
changes)
that
it
induced
in
rats
during
a
2­
year
chronic
feeding
study
(
Walker
et
al.,
1969).
Liver
toxicity
was
noticed
only
at
the
1.0
and
10
ppm
diet
groups
in
female
rats.
The
NOAEL
dietary
concentration
of
0.1
ppm
dieldrin
served
as
the
basis
for
the
RfD
derivation,
after
being
converted
to
a
dose
of
0.005
mg
dieldrin/
kg
bw/
day
utilizing
the
authors'
assumptions
on
dietary
intake
(
which
also
comport
with
the
1
ppm
=
0.05
mg/
kg
bw/
day
conversion
factor
of
Lehman
[
1959]
for
food
consumption
in
rats).

The
RfD
for
dieldrin
is
calculated
as
follows:

RfD
=
=
5E­
5
mg/
kg
bw/
day
0
005
.
mg
/
kg
bw
/
day
100
where:

0.005
mg/
kg
bw/
day
=
NOAEL,
based
on
liver
toxicity
in
rats
exposed
to
dieldrin
in
the
diet
for
2
years
100
=
uncertainty
factor;
this
composite
uncertainty
factor
was
chosen
in
accordance
with
EPA
or
NAS/
OW
guidelines
in
which
uncertainty
factors
of
10
each
were
applied
to
extrapolate
from
rats
to
humans
and
to
account
for
uncertainty
in
the
range
of
human
sensitivity
(
i.
e.,
to
protect
sensitive
human
subpopulations

8.1.2
Reference
Concentration
(
RfC)
Determination
No
human
or
animal
studies
were
identified
that
would
currently
support
the
derivation
of
RfC
values
for
either
aldrin
or
dieldrin.

8.2
Dose­
Response
for
Cancer
Effects
8.2.1
Choice
of
Study/
Data
With
Rationale
and
Justification
Aldrin
Three
marginally
adequate­
to­
adequate
long­
term
carcinogenicity
bioassays
of
aldrin
have
been
conducted
using
B6C3F
1,
C3HeB/
Fe,
and
C3H
mice.
Based
on
these
studies,
there
is
sufficient
evidence
that
aldrin
is
carcinogenic
for
mice.
Dietary
administration
of
aldrin
induced
8­
4
Aldrin/
Dieldrin
 
February
2003
statistically
significant
increases
in
hepatocellular
carcinomas
in
male
B6C3F
1
mice
(
p
<
0.001)
(
NCI,
1978),
hepatomas
in
combined
male
plus
female
C3HeB/
Fe
mice
(
p
<
0.001)
(
Davis
and
Fitzhugh,
1962),
and
hepatomas
in
the
combined
sexes
of
C3H
mice
(
p
<
0.05)
(
Davis,
1965).
Reevaluation
of
the
hepatomas
observed
in
the
latter
two
studies
indicated
that
most
were
actually
hepatocellular
carcinomas
(
Epstein,
1975).

Dietary
administration
of
aldrin
was
reported
to
increase
the
combined
incidences
of
follicular
cell
adenomas
and
carcinomas
of
the
thyroid
in
both
male
and
female
Osborne
Mendel
rats;
however,
the
increase
was
not
dose­
related
and
was
significant
(
p
=
0.001)
only
at
the
low
dose.
This
increase
was
not
considered
to
be
treatment
related.
Although
the
study
authors
concluded
that
aldrin
was
not
carcinogenic
to
rats
(
NCI,
1978),
these
data
(
in
conjunction
with
the
adrenal
cortex
tumors
observed
in
low­
dose
female
rats)
have
been
subsequently
considered
equivocal
or
suggestive
evidence
of
carcinogenicity
in
rats
(
Griesemer
and
Cueto,
1980;
Haseman
et
al.,
1987;
USEPA,
1987).

Based
on
the
available
data,
IARC
(
1987)
concluded
that
there
was
limited
evidence
for
the
carcinogenicity
of
aldrin
in
animals
and
inadequate
evidence
in
humans.
IARC's
conclusion
with
respect
to
animal
carcinogenicity
was
based
on
the
occurrence
of
malignant
liver
neoplasms
in
mice,
since
one
study
using
rats
could
not
clearly
associate
the
occurrence
of
thyroid
tumors
with
aldrin
treatment,
three
additional
studies
using
rats
gave
negative
results,
and
another
rat
study
was
judged
to
be
inadequate.
Consequently,
IARC
classified
aldrin
as
a
Group
3
chemical,
a
possible
human
carcinogen.

Applying
the
criteria
described
in
EPA's
guidelines
for
assessment
of
carcinogenic
risk
(
USEPA,
1986),
aldrin
may
be
classified
in
Group
B2:
probable
human
carcinogen.
This
category
includes
agents
for
which
there
is
inadequate
evidence
of
carcinogenicity
in
human
studies
and
sufficient
evidence
of
carcinogenicity
in
animal
studies.
Under
the
more
recent
Proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
USEPA,
1996/
1999),
aldrin
would
probably
be
categorized
as
"
likely"
to
produce
cancer
in
humans
by
the
oral
route
of
exposure,
while
its
carcinogenic
potential
via
other
routes
of
exposure
would
merit
a
classification
of
"
cannot
be
determined
due
to
inadequate
data."

From
the
several
carcinogenicity
studies
that
have
provided
evidence
that
aldrin
is
carcinogenic
in
mice,
three
data
sets
have
been
deemed
adequate
for
quantitative
risk
estimation
(
USEPA,
1987):
those
for
both
male
and
female
C3H
mice
in
the
Davis
(
1965)
study,
as
reevaluated
by
Reuber
and
cited
in
Epstein
(
1975);
and
that
for
male
B6C3F
1
mice
in
the
NCI
(
1978)
bioassay.
Utilizing
the
linearized
multistage
model,
the
USEPA
(
1987)
performed
potency
estimates
for
each
of
these
data
sets
after
interspecies
dose
conversion;
they
ranged
from
12
to
23
(
mg/
kg
bw/
day)­
1.
Their
geometric
mean,
(
q1*)
=
17
(
mg/
kg
bw/
day)­
1,
was
estimated
as
the
cancer
potency
of
aldrin
for
the
general
population.

Using
this
cancer
potency
estimate
and
assuming
that
a
70­
kg
human
adult
consumes
2
liters
of
water
a
day
over
a
70­
year
lifespan,
the
linearized
multistage
model
yields
a
drinking
water
unit
risk
of
4.9
E­
4
per
:
g/
L.
This
in
turn
can
be
used
to
estimate
that
concentrations
of
8­
5
Aldrin/
Dieldrin
 
February
2003
0.2,
0.02,
and
0.002
:
g/
liter
of
aldrin
may
result
in
excess
cancer
risks
of
10­
4,
10­
5,
and
10­
6,
respectively.

The
linearized
multistage
model
is
only
one
of
several
that
can
be
used
for
estimating
carcinogenic
risk.
From
the
three
aldrin
data
sets
that
were
identified
in
the
USEPA
(
1987)
report
as
being
suitable
for
quantitative
cancer
risk
estimation,
it
was
determined
that
one
was
also
suitable
for
determining
slope
estimates
using
the
probit,
logit,
Weibull,
and
gamma­
multihit
models.
Each
model
utilizes
a
different
set
of
assumptions
in
order
to
extrapolate
from
observed
experimental
data
to
predicted
cancer
risks
at
the
low
doses
more
characteristic
of
human
exposure
scenarios.
Based
on
current
limitations
in
the
understanding
of
biological
mechanisms
relevant
to
carcinogenesis,
as
well
as
in
the
availability
of
mechanistic
data
for
most
chemicals,
the
relative
accuracy
of
these
models
cannot
generally
be
predicted.
The
drinking
water
levels
of
aldrin
estimated
by
each
of
these
models
(
at
the
upper
95%
confidence
limit)
to
be
associated
with
an
excess
cancer
risk
of
one
per
1,000,000
persons
exposed
(
i.
e.,
an
excess
risk
of
10­
6)
were
0.00206
:
g/
L
(
multistage
model),
0.00356
:
g/
L
(
probit
model),
0.00376
:
g/
L
(
logit
model),
0.00356
:
g/
L
(
Weibull
model),
and
0.00310
:
g/
L
(
multihit
model)
(
USEPA,
1992).

Dieldrin
Applying
the
criteria
described
in
EPA's
final
guidelines
for
assessment
of
carcinogenic
risk
(
USEPA,
1986),
dieldrin
also
may
be
classified
in
Group
B2,
probable
human
carcinogen.
Again,
under
the
more
recent
Proposed
Guidelines
for
Carcinogen
Risk
Assessment
(
USEPA,
1996/
1999),
dieldrin
would
probably
be
categorized
as
"
likely"
to
produce
cancer
in
humans
by
the
oral
route
of
exposure,
while
its
carcinogenic
potential
via
other
routes
of
exposure
would
merit
a
classification
of
"
cannot
be
determined
due
to
inadequate
data."
IARC
(
1982)
has
concluded
that
there
is
limited
evidence
for
dieldrin's
carcinogenicity
in
laboratory
animals.

Evidence
reported
in
a
number
of
carcinogenicity
studies
indicates
that
dieldrin
is
carcinogenic
to
several
strains
of
mice
(
Davis
and
Fitzhugh,
1962;
Davis,
1965;
Walker
et
al.,
1972;
Thorpe
and
Walker,
1973;
NCI,
1978;
Tennekes
et
al.,
1981;
Meierhenry
et
al.,
1983).
Thirteen
sex
and
strain­
specific
data
sets
from
these
studies
were
judged
adequate
for
quantitative
risk
estimation;
for
each
of
them,
the
USEPA
generated
potency
estimates
utilizing
the
linearized
multistage
model
(
USEPA,
1987).
These
estimates
ranged
from
7
to
55
(
mg/
kg
bw/
day)­
1,
with
the
geometric
mean
of
q
1*
=
16
(
mg/
kg
bw/
day)­
1
taken
as
the
estimated
potency
of
dieldrin
for
the
general
population.

Using
this
q
1*
value
and
assuming
that
a
70­
kg
human
adult
consumes
2
liters
of
water
a
day
over
a
70­
year
lifespan,
the
linearized
multistage
model
estimates
a
drinking
water
unit
risk
of
4.6
E­
4
per
:
g/
L.
Therefore,
excess
cancer
risk
levels
of
10­
4
,
10­
5,
and
10­
6
would
be
estimated
to
result
from
drinking
water
concentrations
of
approximately
0.2,
0.02,
and
0.002
:
g
dieldrin
per
liter,
respectively.

As
noted
previously,
the
linearized
multistage
model
is
only
one
of
several
that
can
be
used
for
estimating
carcinogenic
risk.
From
the
13
dieldrin
data
sets
data
that
were
identified
in
the
USEPA
(
1987)
report
as
being
suitable
for
quantitative
cancer
risk
estimation,
it
was
8­
6
Aldrin/
Dieldrin
 
February
2003
determined
that
5
were
also
suitable
for
determining
slope
estimates
using
the
probit,
logit,
Weibull,
and
gamma­
multihit
models.
Again,
each
model
utilizes
a
different
set
of
assumptions
in
order
to
extrapolate
from
observed
experimental
data
to
predicted
cancer
risks
at
the
low
doses
more
characteristic
of
human
exposure
scenarios.
Based
on
current
limitations
in
the
understanding
of
biological
mechanisms
relevant
to
carcinogenesis,
as
well
as
in
the
availability
of
mechanistic
data
for
most
chemicals,
the
relative
accuracy
of
these
models
cannot
generally
be
predicted.
The
drinking
water
unit
risks
(
those
estimated
for
a
70
kg
human
drinking,
over
a
the
course
of
a
lifetime,
2
L/
day
of
water
containing
1
:
g/
L
of
dieldrin)
estimated
by
each
of
these
models
(
at
the
upper
95%
confidence
limit)
have
been
reported
as
4.78
×
10­
4
(
multistage
model),
7.7
×
10­
12
(
probit
model),
5.09
×
10­
6
(
logit
model),
1.13
×
10­
4
(
Weibull
model),
and
5.68
×
10­
4
(
multihit
model)
(
USEPA,
1988).
8­
7
Aldrin/
Dieldrin
 
February
2003
References
ATSDR.
2000.
Agency
for
Toxic
Substances
and
Disease
Registry.
Toxicological
profile
for
aldrin/
dieldrin
(
Update).
Draft
for
public
comment.
Atlanta,
GA:
U.
S.
Dept.
of
Health
and
Human
Services,
Public
Health
Service,
ATSDR.

Davis,
K.
J.
1965.
Pathology
report
on
mice
fed
aldrin,
dieldrin,
heptachlor
or
heptachlor
epoxide
for
two
years.
Internal
FDA
memorandum
to
Dr.
A.
J.
Lehman.
FDA.
July
19.
(
as
cited
in
USEPA,
1992).

Davis,
K.
J.
and
O.
G.
Fitzhugh.
1962.
Tumorigenic
potential
of
aldrin
and
dieldrin
for
mice.
Toxicol.
Appl.
Pharmacol.
4:
187­
189
(
as
cited
in
USEPA,
1992).

Deichmann,
W.
B.,
W.
E.
MacDonald,
E.
Blum,
M.
Bevilacqua,
J.
Radomski,
M.
Keplinger,
and
M.
Balkus.
1970.
Tumorigenicity
of
aldrin,
dieldrin
and
endrin
in
the
albino
rat.
Ind.
Med.
Surg.
39(
10):
426­
434
(
as
cited
in
USEPA,
1992).

Epstein,
S.
S.
1975.
The
carcinogenicity
of
dieldrin.
Part
1.
Sci.
Total
Environ.
4:
1­
52
(
as
cited
in
USEPA
1993b).

Fitzhugh,
O.
G.,
A.
A.
Nelson,
and
M.
L.
Quaife.
1964.
Chronic
oral
toxicity
of
aldrin
and
dieldrin
in
rats
and
dogs.
Food
Cosmet.
Toxicol.
2:
551­
562.

Griesemer,
R.
A.
and
C.
Cueto,
Jr.
1980.
Toward
a
classification
scheme
for
degrees
of
experimental
evidence
for
the
carcinogenicity
of
chemicals
for
animals.
In:
Montesano,
R.,
H.
Bartsch,
and
L.
Tomatis,
eds.
Molecular
and
cellular
aspects
of
carcinogen
screening
tests.
IARC
Scientific
Publications
No.
27.
Lyon,
France:
International
Agency
for
Research
on
Cancer,
pp.
259­
281.

Haseman,
J.
K.,
J.
E.
Huff,
E.
Zeiger,
and
E.
E.
McConnell.
1987.
Comparative
results
of
327
chemical
carcinogenicity
studies.
Environ.
Health
Perspect.
74:
229­
235.

IARC.
1987.
International
Agency
for
Research
on
Cancer.
Evaluation
of
the
carcinogenic
risk
of
chemicals
to
humans.
Overall
evaluations
of
carcinogenicity.
Suppl.
7:
88­
89.

IARC.
1982.
International
Agency
for
Research
on
Cancer.
IARC
monographs
on
the
evaluation
of
the
carcinogenic
risk
of
chemicals
to
humans.
Chemicals,
industry
process
and
industries
associated
with
cancer
in
humans.
IARC
Monographs.
Vols.
1­
29,
Supplement
4.
Geneva,
Switzerland:
World
Health
Organization
(
as
cited
in
USEPA,
1988).

IPCS.
1989.
International
Programme
on
Chemical
Safety.
Aldrin
and
dieldrin.
Environmental
health
criteria
91.
Geneva,
Switzerland:
World
Health
Organization,
IPCS.
8­
8
Aldrin/
Dieldrin
 
February
2003
Lehman,
A.
1959.
Appraisal
of
the
safety
of
chemicals
in
foods,
drugs
and
cosmetics.
Association
of
Food
and
Drug
Officials
of
the
United
States
(
as
cited
in
USEPA,
1992
and
USEPA,
1988).

Meierhenry,
E.
F.,
B.
H.
Reuber,
M.
E.
Gershwin,
L.
S.
Hsieh,
and
S.
W.
French.
1983.
Deildrin­
induced
mallory
bodies
in
hepatic
tumors
of
mice
of
different
strains.
Hepatology
3:
90­
95
(
as
cited
in
USEPA,
1993b).

NCI.
1978.
National
Cancer
Institute.
Bioassays
of
aldrin
and
dieldrin
for
possible
carcinogenicity.
DHEW
Publication
No.
(
NIH)
78­
821
and
78­
822.
National
Cancer
Institute
Carcinogenesis
Technical
Report
Series,
No.
21
and
22.
NCI­
CG­
TR­
21,
NCI­
CG­
TR­
22
(
as
cited
in
ATSDR,
2000;
IPCS,
1989;
USEPA,
1993a,
1993b,
1992,
1988,
1987).

Tennekes,
H.
A.,
A.
S.
Wright,
K.
M.
Dix,
and
J.
H.
Koeman.
1981.
Effects
of
dieldrin,
diet,
and
bedding
on
enzyme
function
and
tumor
incidence
in
livers
of
male
CF­
1
mice.
Cancer
Res.
41:
3615­
3620
(
as
cited
in
USEPA,
1993b).

Thorpe,
E.
and
A.
I.
T.
Walker.
1973.
The
toxicology
of
dieldrin
(
HEOD).
Part
II.
Comparative
long­
term
oral
toxicology
studies
in
mice
with
dieldrin,
DDT,
phenobarbitone,
beta­
BHC
and
gamma­
BHC.
Food
Cosmet.
Toxicol.
11:
433­
441
(
as
cited
in
USEPA,
1993b).

Treon,
J.
F.
and
F.
P.
Cleveland.
1955.
Toxicity
of
certain
chlorinated
hydrocarbon
insecticides
for
laboratory
animals,
with
special
reference
to
aldrin
and
dieldrin.
Agric.
Food
Chem.
3:
402­
408
(
as
cited
in
USEPA,
1987,
1992,
1993a
and
1993b).

USEPA.
1996/
1999.
U.
S.
Environmental
Protection
Agency.
Proposed
Cancer
Guidelines.
Available
on
the
Internet
at
http://
www.
epa.
gov/
ORD/
WebPubs/
carcinogen
/
carcin.
pdf.

USEPA.
1993a.
U.
S.
Environmental
Protection
Agency.
IRIS
document
for
aldrin.
Available
on
the
Internet
at
http://
www.
epa.
gov/
ngispgm3/
iris/
index.
html.

USEPA.
1993b.
U.
S.
Environmental
Protection
Agency.
IRIS
document
for
dieldrin.
Available
on
the
Internet
at
http://
www.
epa.
gov/
ngispgm3/
iris/
index.
html.

USEPA.
1992.
U.
S.
Environmental
Protection
Agency.
Aldrin
drinking
water
health
advisory.
Washington,
DC:
USEPA,
Office
of
Water.

USEPA.
1988.
U.
S.
Environmental
Protection
Agency.
Dieldrin
health
advisory.
Washington,
DC:
USEPA,
Office
of
Water.

USEPA.
1987.
U.
S.
Environmental
Protection
Agency.
Carcinogenicity
assessment
of
aldrin
and
dieldrin.
EPA/
600/
6­
87­
006.
Washington
DC:
USEPA,
Office
of
Health
and
Environmental
Assessment,
Carcinogen
Assessment
Group.
8­
9
Aldrin/
Dieldrin
 
February
2003
USEPA.
1986.
U.
S.
Environmental
Protection
Agency.
Guidelines
for
carcinogen
risk
assessment.
Fed.
Reg.
51(
185):
33992­
34003.
September
24
(
as
cited
in
USEPA,
1992,
and
USEPA,
1988).

Walker,
A.
I.
T.,
E.
Thorpe,
and
D.
E.
Stevenson.
1972.
The
toxicology
of
dieldrin
(
HEOD).
I.
Long­
term
oral
toxicity
studies
in
mice.
Food
Cosmet.
Toxicol.
11:
415­
432
(
as
cited
in
USEPA,
1988
and
USEPA,
1993b).

Walker,
A.
I.
T.,
D.
E.
Stevenson,
J.
Robinson,
E.
Thorpe,
and
M.
Roberts.
1969.
The
toxicology
and
pharmacodynamics
of
dieldrin
(
HEOD):
Two­
year
oral
exposures
of
rats
and
dogs.
Toxicol.
Appl.
Pharmacol.
15:
345­
373.
9­
1
Aldrin/
Dieldrin
 
February
2003
9.0
REGULATORY
DETERMINATION
AND
CHARACTERIZATION
OF
RISK
FROM
DRINKING
WATER
9.1
Regulatory
Determination
for
Chemicals
on
the
CCL
The
Safe
Drinking
Water
Act
(
SDWA),
as
amended
in
1996,
required
the
Environmental
Protection
Agency
(
EPA)
to
establish
a
list
of
contaminants
to
aid
the
Agency
in
regulatory
priority
setting
for
the
drinking
water
program.
EPA
published
a
draft
of
the
first
Contaminant
Candidate
List
(
CCL)
on
October
6,
1997
(
62
FR
52193,
USEPA,
1997).
After
review
of
and
response
to
comments,
the
final
CCL
was
published
on
March
2,
1998
(
63FR
10273,
USEPA,
1998).
The
CCL
grouped
contaminants
into
three
major
categories
as
follows:

Regulatory
Determination
Priorities
­
Chemicals
or
microbes
with
adequate
data
to
support
a
regulatory
determination,

Research
Priorities
­
Chemicals
or
microbes
requiring
research
for
health
effects,
analytical
methods,
and/
or
treatment
technologies,

Occurrence
Priorities
­
Chemicals
or
microbes
requiring
additional
data
on
occurrence
in
drinking
water.

The
March
2,
1998,
CCL
included
1
microbe
and
19
chemicals
in
the
regulatory
determination
priority
category.
More
detailed
assessments
of
the
completeness
of
the
health,
treatment,
occurrence
and
analytical
method
data
led
to
a
subsequent
reduction
of
the
regulatory
determination
priority
chemicals
to
a
list
of
12
(
1
microbe
and
11
chemicals),
which
was
distributed
to
stakeholders
in
November
1999.

SDWA
requires
EPA
to
make
regulatory
determinations
for
no
fewer
than
five
contaminants
in
the
regulatory
determination
priority
category
by
August
2001.
In
cases
where
the
Agency
determines
that
a
regulation
is
necessary,
the
regulation
should
be
proposed
by
August
2003
and
promulgated
by
February
2005.
The
Agency
is
given
the
freedom
to
also
determine
that
there
is
no
need
for
a
regulation
if
a
chemical
on
the
CCL
fails
to
meet
one
of
three
criteria
established
by
SDWA
and
described
in
Section
9.1.1.

9.1.1
Criteria
for
Regulatory
Determination
These
are
the
criteria
used
to
determine
whether
or
not
to
regulate
a
chemical
on
the
CCL:

The
contaminant
may
have
an
adverse
effect
on
the
health
of
persons,

The
contaminant
is
known
to
occur
or
there
is
a
substantial
likelihood
that
the
contaminant
will
occur
in
public
water
systems
with
a
frequency
and
at
levels
of
public
health
concern,
9­
2
Aldrin/
Dieldrin
 
February
2003
In
the
sole
judgment
of
the
Administrator,
regulation
of
such
contaminant
presents
a
meaningful
opportunity
for
health
risk
reduction
for
persons
served
by
public
water
systems.

The
findings
for
all
criteria
are
used
in
making
a
determination
to
regulate
a
contaminant.
As
required
by
SDWA,
a
decision
to
regulate
commits
the
EPA
to
publication
of
a
Maximum
Contaminant
Level
Goal
(
MCLG)
and
promulgation
of
a
National
Primary
Drinking
Water
Regulation
(
NPDWR)
for
that
contaminant.
The
Agency
may
determine
that
there
is
no
need
for
a
regulation
when
a
contaminant
fails
to
meet
one
of
the
criteria.
A
decision
not
to
regulate
a
contaminant
is
considered
a
final
Agency
action
and
is
subject
to
judicial
review.
The
Agency
can
choose
to
publish
a
Health
Advisory
(
a
nonregulatory
action)
or
other
guidance
for
any
contaminant
on
the
CCL
independent
of
its
regulatory
determination.

9.1.2
National
Drinking
Water
Advisory
Council
Recommendations
In
March
2000,
the
EPA
convened
a
Working
Group
under
the
National
Drinking
Water
Advisory
Council
(
NDWAC)
to
help
develop
an
approach
for
making
regulatory
determinations.
The
Working
Group
developed
a
protocol
for
analyzing
and
presenting
the
available
scientific
data
and
recommended
methods
to
identify
and
document
the
rationale
supporting
a
regulatory
determination
decision.
The
NDWAC
Working
Group
report
was
presented
to
and
accepted
by
the
entire
NDWAC
in
July
2000.

Because
of
the
intrinsic
differences
between
microbial
and
chemical
contaminants,
the
Working
Group
developed
separate
but
similar
protocols
for
microorganisms
and
chemicals.
The
approach
for
chemicals
was
based
on
an
assessment
of
the
impact
of
acute,
chronic,
and
lifetime
exposures,
as
well
as
a
risk
assessment
that
includes
evaluation
of
occurrence,
fate,
and
dose­
response.
The
NDWAC
Protocol
for
chemicals
is
a
semi­
quantitative
tool
for
addressing
each
of
the
three
CCL
criteria.
The
NDWAC
requested
that
the
Agency
use
good
judgement
in
balancing
the
many
factors
that
need
to
be
considered
in
making
a
regulatory
determination.

The
EPA
modified
the
semi­
quantitative
NDWAC
suggestions
for
evaluating
chemicals
against
the
regulatory
determination
criteria
and
applied
them
in
decision
making.
The
quantitative
and
qualitative
factors
for
aldrin
and
dieldrin
that
were
considered
for
each
of
the
three
criteria
are
presented
in
the
sections
that
follow.

9.2
Health
Effects
The
first
criterion
asks
if
the
contaminant
may
have
an
adverse
effect
on
the
health
of
persons.
Because
all
chemicals
have
adverse
effects
at
some
level
of
exposure,
the
challenge
is
to
define
the
dose
at
which
adverse
health
effects
are
likely
to
occur,
and
estimate
a
dose
at
which
adverse
health
effects
are
either
not
likely
to
occur
(
threshold
toxicant),
or
have
a
low
probability
for
occurrence
(
non­
threshold
toxicant).
The
key
elements
that
must
be
considered
in
evaluating
the
first
criterion
are
the
mode(
s)
of
action,
the
critical
effect(
s),
the
dose­
response
for
critical
effect(
s),
the
RfD
for
threshold
effects,
and
the
slope
factor
for
non­
threshold
effects.
9­
3
Aldrin/
Dieldrin
 
February
2003
A
description
of
the
health
effects
associated
with
exposures
to
aldrin
or
dieldrin
is
presented
in
Chapter
7
of
this
document,
and
is
summarized
below
in
Section
9.2.2.
Chapter
8
and
Section
9.2.3
present
dose­
response
information.

9.2.1
Health
Criterion
Conclusions
The
data
available
on
aldrin
and
dieldrin
demonstrate
the
capacity
of
both
chemicals
to
cause
a
variety
of
adverse
systemic,
neurological,
reproductive/
developmental,
immunological,
genotoxic,
and/
or
tumorigenic
effects
in
humans,
animals,
or
both.
While
some
of
these
noncancer
effects
are
observed
only
at
moderate
to
relatively
high
doses,
others
have
been
observed
to
occur
at
doses
below
0.1
mg/
kg
bw/
day.
The
current
oral
RfDs
for
aldrin
and
dieldrin
are
3
×
10­
5
and
5
×
10­
5
mg/
kg
bw/
day
based
on
hepatic
effects.

Both
compounds
have
been
convincingly
demonstrated
to
be
hepatocarcinogenic
in
several
strains
of
mice
in
multiple
bioassays,
although
they
are
apparently
not
carcinogenic
to
rats
and
have
not
been
convincingly
associated
with
human
cancer
in
any
of
several
large
epidemiology
studies.
Based
on
the
mouse
studies
and
using
the
linear
multistage
model,
the
cancer
potency
for
aldrin
is
17
(
mg/
kg/
day)­
1,
and
that
for
dieldrin,
16
(
mg/
kg/
day)­
1.
For
both
compounds,
a
drinking
water
concentration
of
0.002
:/
L
would
lead
to
an
estimated
lifetime
excess
cancer
risk
of
10­
6.

9.2.2
Hazard
Characterization
and
Mode
of
Action
Implications
Following
acute
exposure
to
high
doses,
the
primary
adverse
health
effects
of
aldrin
and
dieldrin
in
humans
are
those
resulting
from
neurotoxicity
to
the
central
nervous
system,
including
hyperirritability,
convulsions,
and
coma
(
Jager,
1970;
Spiotta,
1951;
ACGIH,
1984).
In
some
cases,
these
may
be
followed
by
cardiovascular
effects,
such
as
tachycardia
and
elevated
blood
pressure
(
Black,
1974).
Under
conditions
of
longer­
term
exposure
to
lower
doses
of
these
compounds,
neurotoxic
symptoms
may
also
include
headache,
dizziness,
general
malaise,
nausea,
vomiting,
and
muscle
twitching
or
myoclonic
jerking
(
Jager,
1970;
ATSDR,
2000a).
Dieldrin
exposure
has
been
linked
to
two
cases
of
immunohemolytic
anemia
(
Hamilton
et
al.,
1978;
Muirhead
et
al.,
1959),
as
has
aldrin/
dieldrin
exposure
to
several
instances
of
aplastic
anemia
(
de
Jong,
1991;
Pick
et
al.,
1965;
ATSDR,
2000a).
However,
at
least
some
of
these
associations
are
problematic,
and
in
any
case,
hematological
or
immunological
(
e.
g.,
dermal
sensitization)
effects
have
not
generally
been
found
in
humans
following
exposure
to
either
compound.

Common
acute
or
subchronic
neurotoxic
effects
observed
in
animals
are
characterized
by
increased
irritability,
salivation,
hyperexcitability,
tremors
followed
by
convulsions,
loss
of
body
weight,
depression,
prostration,
and
death
(
Borgmann
et
al.,
1952;
Walker
et
al.,
1969;
Wagner
and
Greene,
1978;
Woolley
et
al.,
1985;
NCI,
1978;
Casteel,
1993).
These
symptoms
are
similar
to
those
described
above
for
humans
exposed
to
aldrin
or
dieldrin.
Various
manifestations
of
hepatotoxicity
(
elevated
serum
enzyme
levels,
reduced
levels
of
serum
proteins,
hyperplasia,
focal
degeneration,
necrosis,
bile
duct
proliferation,
etc.)
have
been
observed
in
animals
following
subchronic­
to­
chronic
exposure
to
moderate­
to­
high
concentrations
of
aldrin/
dieldrin
9­
4
Aldrin/
Dieldrin
 
February
2003
(
ATSDR,
2000a).
Relatively
low­
dose,
chronic
exposures
to
either
compound
have
been
associated
with
histopathological
liver
changes
in
rat
studies
(
e.
g.,
Fitzhugh
et
al.,
1964;
Walker
et
al.,
1969).
There
is
some
evidence
from
animals
that
aldrin/
dieldrin
exposure
may
either
induce
renal
lesions
or
exacerbate
pre­
existing
nephropathy
(
ATSDR,
2000a;
Fitzhugh
et
al.,
1964;
Harr
et
al.,
1970).

Various
in
vivo
and
in
vitro
studies
have
provided
evidence
that
aldrin
and
dieldrin
may
be
weak
endocrine
disruptors.
Effects
on
male
and
female
hormone
levels
and/
or
receptor
binding,
male
germ
cell
degeneration
and
interstitial
testicular
(
Leydig)
cell
ultrastructure,
estrus
cycle,
and
proliferation
of
endometrial
and
breast
cells
have
been
noted
(
see
Sections
7.3.3
and
7.3.4;
ATSDR,
2000a).
Oral
administration
of
aldrin/
dieldrin
to
maternal
or
paternal
animals
has
produced
somewhat
equivocal
evidence
of
decreased
fertility
(
Dean
et
al.,
1975;
Epstein
et
al.,
1972;
Good
and
Ware,
1969;
Harr
et
al.,
1970;
Virgo
and
Bellward,
1975),
and
intraperitoneal
injection
of
aldrin
has
produced
various
adverse
effects
on
the
male
reproductive
system
(
ATSDR,
2000a).
In
general,
animal
studies
have
provided
only
mixed
evidence
that
exposures
to
aldrin/
dieldrin
at
moderate­
to­
high
levels
can
result
in
adverse
reproductive
or
developmental
effects,
such
as
reduced
fertility
or
litter
size,
reduced
pup
survival,
fetotoxicity,
or
teratogenicity
(
Section
7.2.5).

Immunosuppression
by
dieldrin
has
been
reported
in
a
number
of
mouse
studies:
a
decrease
in
the
antigenic
response
to
the
mouse
hepatitis
virus
3
after
a
single
oral
dose
of
$
18
mg/
kg
bw
(
Krzystyniak
et
al.,
1985);
an
increase
in
the
lethality
of
Plasmodium
berghei
or
Leishmania
tropica
infections
at
dietary
concentrations
as
low
as
1
ppm
(
0.15
mg/
kg
bw/
day)
for
10
weeks
(
Loose,
1982);
and
decreased
tumor
cell
killing
ability
after
dietary
concentrations
as
low
as
1
ppm
(
0.15
mg/
kg
bw/
day)
for
3,
6,
or
18
weeks
(
Loose
et
al.,
1981).

A
number
of
long­
term
bioassay
studies
have
provided
evidence
that
aldrin
and
dieldrin
are
hepatocarcinogens
in
the
mouse
(
Davis
and
Fitzhugh,
1962;
Davis,
1965;
Song
and
Harville,
1964;
NCI,
1978;
MacDonald
et
al.,
1972;
Walker
et
al.,
1972;
Thorpe
and
Walker,
1973;
Tennekes
et
al.,
1982,
1981,
1979;
Meierhenry
et
al.,
1983).
In
one
mouse
study,
dieldrin
was
also
found
to
have
induced
lung,
lymphoid,
and
"
other"
tumors
(
Walker
et
al.,
1972).
In
contrast,
neither
compound
has
been
found
to
induce
liver
tumors
in
various
strains
of
rat
(
Treon
and
Cleveland,
1955;
Song
and
Harville,
1964;
Deichmann
et
al.,
1967,
1970;
Deichmann,
1974;
NCI,
1978;
Fitzhugh
et
al.,
1964;
Walker
et
al.,
1969),
although
a
number
of
these
studies
suffered
from
one
or
more
serious
deficiencies.
The
NCI
(
1978)
rat
study
also
yielded
some
increased
incidences
of
thyroid
follicular
cell
and
adrenal
cortex
adenomas/
carcinomas
following
aldrin
exposure,
which
have
been
considered
either
unrelated
to
treatment
(
NCI,
1978;
USEPA,
1993a),
or
suggestive
of
equivocal
evidence
of
aldrin's
potential
carcinogenicity
in
the
rat
(
Griesemer
and
Cueto,
1980;
Haseman
et
al.,
1987;
USEPA,
1987).

Despite
some
sporadic
statistically
significant
increases
in
rectal
or
liver/
biliary
cancer,
occupational
and
epidemiology
studies
have
failed
to
provide
any
convincing
evidence
for
the
carcinogenicity
of
either
aldrin
or
dieldrin
in
humans
(
Van
Raalte,
1977;
Versteeg
and
Jager,
1973;
de
Jong,
1991;
de
Jong
et
al.,
1997;
Ditraglia
et
al.,
1981;
Brown,
1992;
Amaoteng­
9­
5
Aldrin/
Dieldrin
 
February
2003
Adjepong
et
al.,
1995).
In
fact,
standardized
mortality
ratios
of
exposed
vs.
general
populations
for
both
specific
causes
and
all
causes
of
death
have
generally
been
less
than
1.0.

Not
a
great
deal
is
known
about
the
modes
of
action
that
may
underlie
the
various
toxic
effects
produced
by
exposure
to
aldrin
or
dieldrin.
The
hyperexcitability
associated
with
these
compounds'
neurotoxicity
has
generally
been
thought
to
arise
from
enhancement
of
synaptic
activity
throughout
the
central
nervous
system;
but
whether
this
results
from
facilitated
neurotransmitter
release
at
the
nerve
terminals,
or
from
reducing
the
activity
of
inhibitory
neurotransmitters
within
the
central
nervous
system,
has
been
unclear
(
ATSDR,
2000a).
Mehrota
et
al.
(
1988,
1989)
have
proposed
that
dieldrin
may
act
by
inhibiting
calcium­
dependent
brain
ATPases,
which
would
inhibit
the
cellular
efflux
of
calcium
and
result
in
higher
intracellular
calcium
levels
that
would
promote
neurotransmitter
release.
More
recent
work
provides
significant
evidence
that
aldrin
and
dieldrin's
principal
mode
of
neurotoxic
action
likely
involves
their
role
as
antagonists
for
the
membrane
receptor
for
the
inhibitory
neurotransmitter,
gamma
aminobutyric
acid
(
GABA),
and
blocking
the
influx
of
chloride
ion
through
the
GABA
A
receptor­
ionophore
complex
(
Klaassen,
1996;
Nagata
and
Narahashi,
1994,
1995;
Nagata
et
al.,
1994;
Brannen
et
al.,
1998;
Johns
et
al.,
1998;
Liu
et
al.,
1997,
1998).
Additionally,
at
least
one
in
vitro
study
using
fetal
rat
brain
cells
suggests
that
dieldrin
may
have
an
even
greater
functional
effect
on
dopaminergic
neurons
(
Sanchez­
Ramos
et
al.,
1998).

As
noted
previously,
the
cumulative
evidence
to
date
(
2001)
suggests
that
the
carcinogenic
potential
of
aldrin
and
dieldrin
may
largely
be
limited
to
the
mouse.
The
preponderance
of
evidence
from
the
studies
reviewed
in
this
document
argues
against
a
predominant
role
for
genotoxicity
in
the
mode
of
action
for
these
compounds'
carcinogenicity
(
Sections
7.3.1
and
7.4.2).
This
appears
especially
true
when
considering
the
overwhelmingly
negative
results
for
aldrin
and
dieldrin's
ability
to
induce
gene
point
mutations
(
28
negative
assays,
3
positive
assays).
However,
when
considering
either
direct
DNA
damage
or
chromosome­
related
interactions
(
aberrations,
aneuploidy,
SCEs),
the
assay
results
are
significantly
more
balanced
(
15
negative,
2
most
likely
negative,
11
positive,
4
"
questionably"
positive).

Considering
"
epigenetic"
modes
of
carcinogenic
action,
the
capacity
of
aldrin
and
dieldrin
to
inhibit
various
forms
of
in
vitro
intercellular
communication
in
both
human
and
animal
cells
may
be
significant
with
respect
to
their
in
vivo
effects
on
tumor
production
(
Kurata
et
al.,
1982;
Wade
et
al.,
1986;
Zhong­
Xiang
et
al.,
1986;
Mikalsen
and
Sanner,
1993).
As
discussed
in
Section
7.4.3,
a
number
of
recent
studies
have
provided
suggestive
evidence
that
the
apparent
mouse­
specific
hepatocarcinogenic
effects
of
aldrin/
dieldrin
may
result
from
epigenetic
modes
of
action
that
involve
the
induction
of
intracellular
oxidative
stress
(
via
the
generation
of
reactive
oxygen
species
that
result
in
oxidative
damage
to
DNA,
protein,
and
lipid
macromolecules),
as
well
as
increased
hepatic
DNA
synthesis
(
Kolaja
et
al.,
1995,
1996a,
b,
1998;
Bachowski
et
al.,
1997,
1998;
Stevenson
et
al.,
1995,
1999).
These
effects
have
been
found
to
occur
after
aldrin/
dieldrin
treatment
in
mice,
but
not
in
rats.
After
observing
the
frequency
and
patterns
of
c­
Ha­
ras
protooncogene
mutations
appearing
in
the
DNA
of
glucose­
6­
phosphatase­
deficient
hepatic
lesions
found
in
control
mice,
or
in
those
treated
with
dieldrin
or
phenobarbital,
Bauer­
Haufmann
et
al.
(
1992)
were
able
to
conclude
that
the
increase
in
hepatic
9­
6
Aldrin/
Dieldrin
 
February
2003
lesions
(
and
thus
tumors)
resulting
from
dieldrin
treatment
likely
resulted
primarily
from
promotional,
rather
than
initiation,
events.
It
has
been
postulated
that
aldrin/
dieldrin
induction
of
hepatic
DNA
synthesis
may
also
result
from
the
modulation
of
proto­
oncogene
expression
via
various
transcription
factors
(
Stevenson
et
al.,
1999).

The
available
literature
did
not
provide
direct
evidence
for
any
human
subpopulations
that
would
be
particularly
sensitive
to
the
toxic
effects
of
chronic
aldrin/
dieldrin
exposure,
or
for
which
relevant
toxicokinetics
are
known
to
be
differ
significantly
from
those
for
the
general
population.
Speculatively,
the
fetus
and
very
young
children
might
be
at
increased
risk
from
exposures
to
aldrin/
dieldrin
as
a
result
of
immature
hepatic
detoxification
and
excretion
functions,
as
well
as
developing
target
organ
systems.
Some
support
for
this
is
found
in
a
single
case
study
involving
acute
exposure
to
aldrin
(
Hayes,
1982)
in
which
a
3
year­
old
female
child
died
after
ingesting
approximately
8.2
mg/
kg,
or
roughly
an
order
of
magnitude
less
than
the
estimated
lethal
dose
for
an
adult
male.
Several
mechanistic
studies,
which
describe
the
prenatal
effects
of
aldrin/
dieldrin
on
GABA
receptor
malfunctions
and
on
subsequent
behavioral
impairment,
may
suggest
an
increased
sensitivity
of
children
(
Brannen
et
al.,
1998;
Liu
et
al.,
1998;
Johns
et
al.,
1998;
Castro
et
al.,
1992).
Declining
organ
and
immune
functions
may
also
render
the
elderly
more
susceptible
to
aldrin/
dieldrin
toxicity.
Additionally,
it
is
reasonable
to
expect
that
individuals
with
compromised
liver,
immune,
or
neurological
functions
(
as
a
result
of
disease,
genetic
predisposition,
or
other
toxic
insult)
might
also
display
increased
sensitivity
to
these
compounds.

9.2.3
Dose­
Response
Characterization
and
Implications
in
Risk
Assessment
In
adult
humans,
the
acute
oral
lethal
dose
for
aldrin/
dieldrin
has
been
estimated
at
approximately
70
mg/
kg
bw
(
Jager,
1970;
ATSDR,
2000a),
which
is
about
3
times
the
dose
reported
to
have
induced
convulsions
within
20
minutes
of
ingestion
(
Spiotta,
1951).
Oral
LD
50
values
in
various
animal
species
for
the
two
compounds
have
been
reported
to
range
from
33
to
95
mg/
kg
bw,
and
appear
to
be
affected
by
age
at
the
time
of
exposure.
In
rats,
LD
50
values
were
reported
as
37
mg/
kg
bw
for
young
adults,
25
mg/
kg
bw
for
2­
week­
old
pups,
and
a
somewhat
surprisingly
high
168
mg/
kg
bw
for
newborns
(
Lu
et
al.,
1965).

Meaningful
dose­
response
relationships
have
not
been
adequately
characterized
in
humans
for
any
of
the
toxic
effects
of
aldrin
or
dieldrin.
In
animals,
oral
exposure
to
aldrin/
dieldrin
has
produced
a
variety
of
dose­
dependent
systemic,
neurological,
immunological,
endocrine,
reproductive,
developmental,
genotoxic,
and
tumorigenic
effects
over
a
collective
dose
range
of
at
least
three
orders
of
magnitude
(<
0.05
to
50
mg/
kg
bw),
depending
on
the
specific
endpoint
and
the
duration
of
exposure
(
Sections
7.2
and
7.3)
(
ATSDR,
2000a).
Doseresponse
information
for
some
key
studies
is
summarized
below
in
Table
9­
1.
For
noncancer
effects,
the
USEPA
has
determined
oral
RfDs
for
both
aldrin
and
dieldrin
(
see
Sections
8.1.1.1
and
8.1.1.2)
based
on
the
most
sensitive
relevant
toxic
effects
(
critical
effects)
that
have
been
reported.
For
aldrin,
the
critical
effect
was
liver
toxicity
observed
in
rats
after
chronic
exposure
to
approximately
0.025
mg/
kg
bw/
day,
the
LOAEL
and
lowest
dose
tested
(
Fitzhugh
et
al.,
1964).
This
dose
was
divided
by
a
composite
uncertainty
factor
of
1,000
(
to
account
for
rat­
tohuman
extrapolation,
potentially
sensitive
human
subpopulations,
and
the
use
of
a
LOAEL
rather
9­
7
Aldrin/
Dieldrin
 
February
2003
than
a
NOAEL)
to
yield
an
oral
RfD
for
aldrin
of
3
×
10­
5
mg/
kg
bw/
day.
Similarly
for
dieldrin,
a
chronic
rat
study
NOAEL
for
liver
toxicity
of
approximately
0.005
mg/
kg
bw/
day
(
Walker
et
al.,
1969)
was
divided
by
a
composite
uncertainty
factor
of
100
(
to
account
for
rat­
to­
human
extrapolation
and
potentially
sensitive
human
subpopulations)
to
yield
an
oral
RfD
of
5
×
10­
5
mg/
kg
bw/
day.

Based
on
the
long­
term
mouse
bioassays
discussed
in
Sections
7.2.7
and
7.4.2
to
7.4.4,
the
USEPA
has
classified
both
aldrin
and
dieldrin
as
group
B2
carcinogens
under
the
current
cancer
guidelines
(
USEPA,
1986),
that
is,
as
probable
human
carcinogens
with
little
or
no
evidence
of
carcinogenicity
in
humans,
and
sufficient
evidence
in
animals.
Under
the
USEPA's
proposed
cancer
risk
assessment
guidelines
(
USEPA,
1996/
1999),
the
weight
of
evidence
indicates
that
aldrin
and
dieldrin
could
be
classified
as
rodent
carcinogens
that
are
"
likely
to
be
carcinogenic
to
humans
by
the
oral
route
of
exposure,
but
whose
carcinogenic
potential
by
the
inhalation
and
dermal
routes
of
exposure
cannot
be
determined
because
there
are
inadequate
data
to
perform
an
assessment."
This
characterization
must
be
tempered
by
the
lack
of
evidence
for
significant
human
carcinogenicity
from
epidemiological
studies
and
by
the
general
lack
of
corroborative
evidence
for
carcinogenicity
in
rats.
Mechanistic
studies
performed
in
vitro
and
in
vivo
suggest
that
one
or
more
non­
genotoxic
modes
of
action
may
underlie
or
contribute
to
the
carcinogenic
potential
of
aldrin
and
dieldrin,
but
these
effects
are
not
completely
established,
nor
can
a
role
for
genotoxic
mechanisms
confidently
be
eliminated
based
on
the
available
data.
Based
on
these
considerations,
the
quantitative
cancer
risk
assessments
of
aldrin
and
dieldrin
have
been
conducted
conservatively
using
the
linear­
default
model.
9­
8
Aldrin/
Dieldrin
 
February
2003
Table
9­
1.
Dose­
Response
Information
from
Key
Studies
of
Aldrin
and
Dieldrin
Toxicity
Study
Species
No./
Sex
per
Group
Doses
mg/
kg
bw/
day
Duration
NOAEL
mg/
kg
bw/
day
LOAEL
mg/
kg
bw/
day
Effects
Chronic
Studies
S
Aldrin1
Fitzhugh
et
al.

(
1964)
Rat
Osborne­

Mendel
12
M
12
F
0
0.025
0.1
0.5
2.5
5.0
7.5
2
yr
S
0.5
0.025
2.5
Liver
histopathology
Increased
mortality;
enlarged
livers;
nephritis;
distended
and
hemorrhagic
urinary
bladders
Chronic
Studies
S
Dieldrin1
Walker
et
al.
(
1969)
Rat
Carworth
Farm
"
E"
25
M
25
F
0
0.005
0.05
0.5
2
yr
0.005
0.05
0.05
0.5
Increased
absolute
and
relative
liver
weights
Irritability,
tremors,
convulsions;

CHIRL2
Cancer
Bioassay
Studies
S
Aldrin3
Davis
(
1965)
4
Mouse
C
3
H
100
M
100
F
0
1.5
2
yr
S
1.5
Hepatomas
and
hepatocellular
carcinomas
(
not
tabulated
by
sex)
Table
9­
1
(
continued)

Study
Species
No./
Sex
per
Group
Doses
mg/
kg
bw/
day
Duration
NOAEL
mg/
kg
bw/
day
LOAEL
mg/
kg
bw/
day
Effects
9­
9
Aldrin/
Dieldrin
 
February
2003
NCI
(
1978)
Mouse
B6C3F
1
50
M
50
F
0
0.45
(
F)

0.6
(
M)

0.9
(
F)

1.2
(
M)
80
wk
S
0.6
(
M)
Hepatocellular
carcinoma
(
M);
no
statistically
significant
tumor
increases
were
observed
in
(
F)

NCI
(
1978)
5
Rat
Osborne­

Mendel
50
M
50
F
0
1.5
3
74
wk
(
M)

80
wk
(
F)
S
1.5
?
Suggestive/
equivocal
evidence
of
thyroid
follicular
cell
adenoma
and
carcinoma
(
M/
F)
and
adrenal
cortex
adenoma
(
F)
at
low,
but
not
high,
dose
Cancer
Bioassay
Studies
S
Dieldrin3
Davis
(
1965)
4
Mouse
C
3
H
100
M
100
F
0
1.5
2
yr
S
1.5
Hepatomas
and
hepatocellular
carcinomas
(
not
tabulated
by
sex)
Table
9­
1
(
continued)

Study
Species
No./
Sex
per
Group
Doses
mg/
kg
bw/
day
Duration
NOAEL
mg/
kg
bw/
day
LOAEL
mg/
kg
bw/
day
Effects
9­
10
Aldrin/
Dieldrin
 
February
2003
Walker
et
al.
(
1972)
Mouse
CF
1
125­
300
M
125­
300
F
30
M
30F
0
0.015
0.15
1.5
0
0.188
0.375
0.75
1.5
3
132
wk
128
wk
S
S
0.15
0.375
Hepatocellular
carcinoma
(
F),
no
statistically
significant
tumor
increase
was
observed
at
low
dose;
[
hepatocellular
carcinoma
and
hepatoma
at
high
dose
(
M/
F);

lung
and
lymphoid
tumors
at
low
and
medium
doses
(
F)]

Hepatocellular
carcinomas
and/
or
hepatomas
(
M/
F),
no
statistically
significant
tumor
increases
were
observed
at
the
low
dose
Thorpe
and
Walker
(
1973)
Mouse
CF
1
30­
45
M
30­
45
F
0
1.5
110
wk
S
1.5
Hepatocellular
carcinomas
and
hepatomas
(
M/
F)

NCI
(
1978)
Mouse
B6C3F
1
50
M
50
F
0
0.375
0.75
80
wk
S
0.75
(
M)
Hepatocellular
carcinomas
(
M);

no
statistically
significant
tumor
increases
were
observed
at
the
low
dose
(
M)
or
in
(
F)

Tennekes
et
al.

(
1981)
Mouse
CF
1
139
M
(
total;
252
controls)
0
1.5
110
wk
S
1.5
(
M)
Hepatocellular
carcinomas
and
hepatomas
(
M;
2
experiments
with
different
diets)
Table
9­
1
(
continued)

Study
Species
No./
Sex
per
Group
Doses
mg/
kg
bw/
day
Duration
NOAEL
mg/
kg
bw/
day
LOAEL
mg/
kg
bw/
day
Effects
9­
11
Aldrin/
Dieldrin
 
February
2003
Meierhenry
et
al.

(
1983)
Mouse
C57BL/
6J
Mouse
C3H
Mouse
B6C3F
1
69­
71
M
50
M
62­
76
M
0
1.5
0
1.5
0
1.5
85
wk
85
wk
85
wk
S
S
S
1.5
(
M)

1.5
(
M)

1.5
(
M)
Hepatocellular
carcinomas
(
and
hepatomas
in
C57BL/
6J
and
B6C3F
1
strains)

1
Studies
serving
as
the
principal
basis
for
oral
RfD
determinations.

2
Chlorinated
hydrocarbon
insecticide
rodent
liver.

3
Studies
utilized
in
the
derivation
of
cancer
potency
estimates.

4
As
reevaluated
by
Reuber
and
reported
in
Epstein
(
1975).

5
This
study
was
not
used
for
the
derivation
of
cancer
potency
estimates,
but
is
the
source
of
the
only
data
that
provides
any
evidence
of
aldrin/
dieldrin's
tumorigenic
potential
in
the
rat.
9­
12
Aldrin/
Dieldrin
 
February
2003
This
approach
has
yielded
geometric
mean
cancer
potency
estimates
for
aldrin
and
dieldrin
of
17
and
16
(
mg/
kg
bw/
day)­
1,
respectively
(
Sections
8.2.1.1
and
8.2.1.2).
These
result
in
drinking
water
unit
risks
of
4.9
×
10­
4
per
mg/
L
and
4.6
×
10­
4
per
mg/
L,
respectively.
For
both
compounds,
a
drinking
water
concentration
of
0.002
:
g/
L
would
lead
to
an
estimated
lifetime
excess
cancer
risk
of
10­
6.
This
concentration,
0.002
:
g/
L,
was
selected
as
the
Health
Reference
Level
(
HRL)
for
each
chemical,
and
was
used
in
Chapter
4
to
put
into
context
the
levels
of
aldrin/
dieldrin
detected
in
drinking
water.

9.3
Occurrence
in
Public
Water
Systems
The
second
criterion
asks
if
the
contaminant
is
known
to
occur,
or
if
there
is
a
substantial
likelihood
that
the
contaminant
will
occur,
in
public
water
systems
with
a
frequency
and
at
levels
of
concern
for
public
health.
In
order
to
address
this
question,
the
following
information
was
considered:

°
Monitoring
data
from
public
water
systems
°
Ambient
water
concentrations
and
releases
to
the
environment
°
Environmental
fate
Data
on
the
occurrence
of
aldrin
and
dieldrin
in
public
drinking
water
systems
were
the
most
important
determinants
in
evaluating
the
second
criterion.
EPA
looked
at
the
total
number
of
systems
that
reported
detections
of
aldrin/
dieldrin,
as
well
those
that
reported
concentrations
of
aldrin/
dieldrin
above
an
estimated
drinking
water
health
reference
level
(
HRL).
For
noncarcinogens,
the
estimated
HRL
risk
level
was
calculated
from
the
RfD
assuming
that
20%
of
the
total
exposure
would
come
from
drinking
water.
For
carcinogens,
the
HRL
was
the
10­
6
risk
level.
The
HRLs
are
benchmark
values
that
are
used
in
evaluating
the
occurrence
data
while
the
risk
assessments
for
the
contaminants
are
being
developed.

The
available
monitoring
data,
including
indications
of
whether
or
not
the
contamination
is
a
national
or
a
regional
problem,
are
included
in
Chapter
4
of
this
document
and
are
summarized
below.
Additional
information
on
production,
use,
and
environmental
fate
are
found
in
Chapters
2
and
3.

9.3.1
Occurrence
Criterion
Conclusions
Since
aldrin
and
dieldrin
have
not
been
used
in
this
country
since
1987,
there
should
be
no
new
releases
to
the
overall
environment.
The
analyses
presented
previously
for
aldrin
and
dieldrin
indicate
that
these
chemicals
are
detected
very
infrequently
and
at
very
low
concentrations
in
drinking
water.
Therefore,
it
is
unlikely
that
aldrin
and
dieldrin
will
occur
in
public
water
systems
with
any
significant
frequency
at
levels
of
concern
for
public
health.
9­
13
Aldrin/
Dieldrin
 
February
2003
9.3.2
Monitoring
Data
Drinking
Water
As
more
fully
discussed
in
Chapter
4,
the
analyzed
drinking
water
occurrence
data
for
aldrin
and
dieldrin
were
collected
beginning
in
1993
under
"
Round
2"
of
the
Safe
Drinking
Water
Act's
Unregulated
Contaminant
Monitoring
(
UCM)
Program.
Monitoring
ended
for
small
public
water
systems
(
PWSs)
on
January
8,
1999,
and
for
large
PWSs
on
January
1,
2001.
Round
2
UCM
data
were
collected
from
35
"
primacy
entities,"
which
included
34
states
and
some
Native
American
tribal
systems.
However,
because
the
data
from
some
states
were
incomplete
and/
or
otherwise
biased,
and
because
the
data
were
not
collected
within
a
systematic
or
random
statistical
framework,
the
national
representativeness
of
the
combined
data
set
is
considered
problematic.
In
an
attempt
to
at
least
partially
address
these
concerns,
a
cross­
section
of
state
data
sets
was
constructed
that
provides
a
reasonable
representation
(
although
not
a
truly
"
statistically
representative"
sample)
of
national
occurrence.
This
was
accomplished
by
a
process
of
first
evaluating
the
data
sets
for
completeness,
quality,
and
bias;
after
eliminating
unusable
state
data,
the
remaining
states
were
reevaluated
for
their
pollution
potential
(
from
manufacturing
and
population
density,
and
from
agricultural
activity)
and
their
"
geographic
coverage"
in
relation
to
all
states.
The
result
of
this
process
established
a
"
national
crosssection
of
Round
2
states
(
AK,
AR,
CO,
KY,
ME,
MD,
MA,
MI,
MN,
MS,
NH,
NM,
NC,
ND,
OH,
OK,
OR,
RI,
TX,
and
WA).

It
should
be
noted
that
while
MA
was
included
in
the
Round
2
cross­
section
on
the
basis
of
usable
and
complete
data
for
volatile
organic
compounds
(
VOCs)
and
inorganic
compounds
(
IOCs),
it
was
excluded
from
the
analysis
of
synthetic
organic
compounds
like
aldrin
and
dieldrin
because
of
incomplete
and
abnormal
data
(
atypically
high
percentage
of
detects
in
a
relatively
small
number
of
PWSs).
Therefore,
the
Round
2
cross­
section
(
R2­
X)
data
discussed
here
exclude
that
from
MA
and
are
based
on
the
other
19
states;
selected
summary
statistics
are
shown
in
Table
9­
2.
For
perspective
and
to
provide
a
conservative
"
upper
bound"
analysis
of
aldrin/
dieldrin
occurrence
in
drinking
water,
certain
summary
statistics
and
national
extrapolations
based
on
all
reporting
Round
2
states
(
i.
e.,
"
R2­
ARS"
data)
are
presented
here
and
in
Chapter
4.

The
data
indicate
that
both
compounds
are
only
infrequently
detected
in
PWSs
and
only
at
very
low
concentrations.
Because
the
HRL
(
0.002
:
g/
L)
is
below
all
of
the
states
Minimum
Reporting
Levels
(
MRLs),
any
sample
detect
is
also
greater
than
the
HRL
and
½
HRL
levels;
thus,
summary
occurrence
statistics
are
all
the
same,
whether
based
on
the
MRL,
HRL,
or
½
HRL.
Aldrin
was
detected
in
0.016%
of
the
R2­
X
PWSs
at
concentrations
$
the
HRL,
which
yields
a
national
extrapolation
of
11
PWSs
serving
39,000
people.
Although
excluded
from
the
Round
2
cross­
section,
states
with
positively­
biased
detect
statistics
(
e.
g.,
AL)
nonetheless
represent
real
detections
of
aldrin
in
drinking
water
that
are
not
adequately
accounted
for
by
R2­
X
data
extrapolation.
As
a
consequence,
R2­
X
data
extrapolation
clearly
underestimates
the
national
occurrence
of
aldrin
in
PWSs.
To
provide
a
more
conservative
estimate,
one
which
is
likely
an
overestimate
of
national
occurrence,
R2­
ARS
data
may
be
used
for
extrapolation.
In
9­
14
Aldrin/
Dieldrin
 
February
2003
this
case,
an
R2­
ARS
PWS
detection
rate
of
0.212%
extrapolates
nationally
to
138
PWSs
having
aldrin
concentrations
$
the
HRL,
and
serving
1,052,000
people.

Table
9­
2.
Selected
Summary
Statistics
for
Occurrence
of
Aldrin
and
Dieldrin
in
Drinking
Water
Parameter
Round
2
Cross­
Section
(
19
States)
1
Round
2
Reporting
States2
Aldrin
Total
samples
31,083
41,565
Percent
of
samples
with
detections
0.006%
0.132%

Median
concentration
(
all
samples)
<(
Non­
detect)
<(
Non­
detect)

99th
Percentile
concentration
(
all
samples)
<(
Non­
detect)
<(
Non­
detect)

Median
concentration
(
detections
only)
0.58
:
g/
L
0.18
:
g/
L
99th
Percentile
concentration
(
detections
only)
0.69
:
g/
L
4.40
:
g/
L
Minimum
Reporting
Level
(
MRL)
variable
variable
Draft
Health
Reference
Level
(
HRL)
0.002
:
g/
L
0.002
:
g/
L
Percent
of
PWSs
with
detections
>
MRL
0.016%
0.212%

Percent
of
PWSs
with
detections
>(
1/
2
HRL)
0.016%
0.212%

Percent
of
PWSs
with
detections
>
HRL
0.016%
0.212%

Dieldrin
Total
samples
29,603
40,055
Percent
of
samples
with
detections
0.064%
0.135%

Median
concentration
(
all
samples)
<(
Non­
detect)
<(
Non­
detect)

99th
Percentile
concentration
(
all
samples)
<(
Non­
detect)
<(
Non­
detect)

Median
concentration
(
detections
only)
0.16
:
g/
L
0.42
:
g/
L
99th
Percentile
concentration
(
detections
only)
1.36
:
g/
L
4.40
:
g/
L
Minimum
Reporting
Level
(
MRL)
variable
variable
Draft
Health
Reference
Level
(
HRL)
0.002
:
g/
L
0.002
:
g/
L
Percent
of
PWSs
with
detections
>
MRL
0.093%
0.211%

Percent
of
PWSs
with
detections
>(
1/
2
HRL)
0.093%
0.211%

Percent
of
PWSs
with
detections
>
HRL
0.093%
0.211%

1Based
on
data
from
the
20­
State
Cross
Section,
minus
MA
(
SDWIS/
FED,
UCM
Round
2,
1993).
2Based
on
data
from
all
reporting
states
(
SDWIS/
FED,
UCM
Round
2,
1993).
Source:
Data
taken
from
Tables
4­
2
and
4­
5
in
Section
4.0
of
this
document.
Abbreviations:
HRL
=
Health
Reference
Level;
MRL
=
Minimum
Reporting
Level;
PWS
=
Public
Water
System.
9­
15
Aldrin/
Dieldrin
 
February
2003
Although
only
five
states
(
AL,
MA,
NM,
PA,
TX)
reported
detecting
aldrin
in
any
of
their
PWSs,
their
distribution
is
sufficiently
broad
to
categorize
aldrin's
drinking
water
occurrence
as
national
in
scope,
rather
than
just
regional
or
local.
This
conclusion
is
further
supported
by
the
observations
that
aldrin
has
been
detected
at
NPL
sites
in
at
least
31
states,
and
at
least
in
40
states
at
sites
listed
in
ATSDR's
HazDat
database.
Independent
analysis
of
data
from
the
corn
belt
states
of
Iowa,
Indiana,
and
Illinois
revealed
that
aldrin
was
not
detected
in
any
surface
or
ground
water
PWS
in
Iowa
or
Indiana,
or
in
any
ground
water
PWS
in
Illinois.
It
was,
however,
detected
in
1.8%
of
Illinois'
surface
water
PWSs.

Similarly,
dieldrin
was
detected
in
0.093%
of
the
R2­
X
PWSs
at
concentrations
$
the
HRL,
which
yields
a
national
extrapolation
of
61
PWSs
serving
150,000
people.
As
with
aldrin,
a
more
conservative
estimate
(
a
likely
overestimate)
of
national
dieldrin
occurrence
in
drinking
water
may
be
derived
using
R2­
ARS
data
for
extrapolation.
In
this
case,
an
R2­
ARS
PWS
detection
rate
of
0.211%
extrapolates
nationally
to
137
PWSs
with
dieldrin
concentrations
$
the
HRL,
serving
793,000
people.

Again,
although
only
eight
states
(
AL,
AR,
CT,
MA,
MD,
NC,
PA,
TX)
reported
detecting
dieldrin
in
any
of
their
PWSs,
their
distribution
is
sufficiently
broad
to
categorize
dieldrin's
drinking
water
occurrence
as
national
in
scope,
rather
than
just
regional
or
local.
This
conclusion
is
further
supported
by
the
observation
that
dieldrin
has
been
detected
at
NPL
sites
in
at
least
31
states
and
at
least
in
40
states
at
sites
listed
in
ATSDR's
HazDat
(
ATSDR,
2000b)
database.
Independent
analysis
of
data
from
the
corn
belt
states
of
Iowa,
Indiana,
and
Illinois
revealed
that
dieldrin
was
not
detected
in
any
surface
or
ground
water
PWS
in
Iowa,
or
in
any
ground
water
PWS
in
Illinois
or
Indiana.
It
was,
however,
detected
in
1.8%
of
Illinois'
and
2.1%
of
Indiana's
surface
water
PWSs.

Ambient
Water
In
the
context
of
drinking
water,
"
ambient
water"
may
be
defined
as
source
water
that
exists
in
surface
waters
and
aquifers
before
treatment
(
Chapter
4).
The
U.
S.
Geological
Survey's
(
USGS's)
National
Ambient
Water
Quality
Assessment
(
NAWQA)
Program,
which
began
in
1991,
provides
the
most
comprehensive
and
nationally
representative
data
available
that
describe
ambient
water
quality.
Unfortunately,
aldrin
was
not
selected
as
an
analyte
for
either
the
NAWQA's
ground
water
or
surface
water
studies.
However,
the
NAWQA
did
analyze
for
aldrin
in
aquatic
biota
tissue
and
stream
bed
sediments
at
591
sites
from
20
of
its
59
"
study
units"
(
i.
e.,
significant
watersheds
and
aquifers)
during
the
period
from
1992
to
1995.
While
aldrin
was
not
detected
in
any
aquatic
biota
tissue
samples,
it
was
detected
above
the
Method
Detection
Limit
(
MDL)
of
1
mg/
kg
in
0.4%
of
all
sites
(
urban
=
0.0%,
mixed
land
use
=
0.5%,
agricultural
=
0.6%,
forest­
rangeland
=
0.0%).
Additionally,
a
mid­
1980s
survey
of
community
water
supply
wells
in
Illinois
detected
aldrin
in
only
0.3%
of
the
wells,
using
an
MRL
of
0.004
mg/
L.

In
contrast
to
aldrin,
dieldrin
was
selected
as
an
NAWQA
analyte
for
both
surface
and
ground
water
studies
during
the
first
round
of
intensive
monitoring
(
1991
to
1996),
which
targeted
20
study
unit
watersheds.
Dieldrin
detection
frequencies
at
two
MDLs
(
0.001
mg/
L;
0.01
mg/
L)
were
as
follows
for
stream
surface
waters:
urban
(
3.67%;
1.83%),
integrator
(
3.27%;
9­
16
Aldrin/
Dieldrin
 
February
2003
1.63%),
agricultural
(
6.90%;
3.90%),
and
total
sites
(
4.64%;
2.39%).
For
ground
water
sources,
the
comparable
data
were:
shallow
urban
(
5.65%;
3.32%),
shallow
agricultural
(
0.97%;
0.65%),
major
aquifers
(
0.43%;
0.21%),
and
total
sites
(
1.42%;
0.93%).
As
with
aldrin,
the
NAWQA
program
also
analyzed
for
dieldrin
in
aquatic
biota
tissue
and
stream
bed
sediments
at
591
sites
from
20
of
its
59
"
study
units"
during
the
period
from
1992
to
1995.
It
was
detected
at
levels
above
the
MDL
of
1
mg/
kg
in
13.7%
of
the
sediments
from
all
sites,
and
at
levels
above
the
MDL
of
5
mg/
kg
in
28.6%
of
whole
fish
samples
and
in
6.4%
of
bivalve
samples.
Additionally,
a
1991
to
1992
survey
of
surface
waters
from
the
Mississippi
River
and
six
of
its
tributaries
that
drain
the
corn
belt
reported
8%
of
all
samples
and
71%
of
all
sites
registered
detections
above
the
MRL
of
0.02
mg/
L.

9.3.3
Use
and
Fate
Data
Both
aldrin
and
dieldrin
are
SOC
pesticides
that
were
at
one
time
extensively
used
in
a
wide
variety
of
agricultural
and
residential/
urban
pest­
control
applications
(
Chapters
2
and
4).
They
were
manufactured
and
distributed
in
the
United
States
by
the
Shell
Chemical
Company
until
1974.
From
1974
through
1985
(
except
1979
to
1980),
Shell
International
(
Holland)
imported
lesser
amounts
(
e.
g.,
1
to
1.5
million
lb/
year
from
1981
to
1985).
Importation
information
for
dieldrin
was
not
available.
In
1972,
the
USEPA
cancelled
all
but
three
specific
uses
of
these
compounds
(
subsurface
ground
insertion
for
termite
control,
dipping
of
non­
food
plant
roots
and
tops,
and
moth­
proofing
in
manufacturing
processes
using
completely
closed
systems).
This
decision
was
finalized
in
1974,
and
by
1987
these
remaining
uses
were
voluntarily
cancelled
by
the
manufacturer.

Use
of
aldrin
in
the
U.
S.
peaked
at
19,000,000
lbs
in
1966,
decreasing
to
10,500,000
lbs
by
1970;
during
the
same
period,
dieldrin
use
declined
from
1,000,000
to
670,000
lbs
(
ATSDR,
2000a).
By
the
time
the
Toxic
Release
Inventory
(
TRI)
was
mandated
in
1986
by
the
Emergency
Planning
and
Community
Right­
to­
Know
Act
(
EPCRA)
and
then
subsequently
implemented,
the
manufacture,
import,
and
use
of
aldrin/
dieldrin
had
been
cancelled.
The
EPCRA
mandates
that
facilities
with
more
than
10
full­
time
employees
that
manufacture/
import
more
than
25,000
lbs,
or
use
more
than
10,000
lbs,
of
a
TRI
chemical
are
required
to
report
the
lb/
year
of
the
chemical
that
were
released
to
the
environment,
both
on­
site
and
off­
site.
It
was
not
until
1995
that
hazardous
waste
treatment
and
disposal
facilities
were
added
to
the
list
of
those
required
to
report
TRI
data.
In
1998,
the
first
year
for
which
this
requirement
became
effective,
hazardous
waste
facilities
in
three
states
(
AR,
MI,
TX)
reported
releases
of
aldrin
totaling
25,622
lbs.
No
such
releases
of
dieldrin
were
reported.

The
environmental
fate
of
aldrin
and
dieldrin
is
extensively
summarized
in
Chapter
3.
Briefly,
under
most
environmental
conditions,
aldrin
is
largely
converted
biologically
or
abiotically
to
dieldrin,
which
is
significantly
more
environmentally
stable.
Most
of
these
compounds
are
released
to
the
environment
via
the
soil,
where
relatively
high
log
K
ow
and
K
oc
are
indicative
of
their
low
water
solubility
and
strong
affinity
for
adsorption
to
soil.
Over
time,
significant
quantities
may
volatilize
to
the
atmosphere
or
be
carried
aloft
by
wind­
born
particles,
where
they
are
subject
to
certain
photodegradation
processes
and/
or
subsequent
"
washout"
in
rainfall.
Because
of
their
low
water
solubilities
and
strong
soil
adsorption
tendencies,
aldrin
and
9­
17
Aldrin/
Dieldrin
 
February
2003
dieldrin
slowly
migrate
downward
through
the
soil
or
enter
surface
or
ground
water.
Most
aldrin/
dieldrin
found
in
surface
water
is
thought
to
result
from
particulate
surface
run­
off
(
the
compounds
being
bound
to
soil
particles).
In
summary,
these
characteristics
will
tend
to
maintain
relatively
low
levels
of
water
contamination
over
relatively
prolonged
periods
of
time.

Obviously,
neither
compound
is
used
as
a
drinking
water
treatment
chemical,
nor
is
either
likely
to
be
a
leachate
from
drinking
water
contact
surfaces.
However,
it
is
not
unreasonable
to
expect
that
they
may
co­
occur
in
drinking
water
with
each
other,
as
well
as
with
certain
other
persistent
pesticides;
in
such
cases,
additive
or
synergistic
toxic
effects
may
be
possible.

9.4
Risk
Reduction
The
third
criterion
asks
if,
in
the
sole
judgment
of
the
Administrator,
regulation
presents
a
meaningful
opportunity
for
health
risk
reduction
for
persons
served
by
public
water
systems.
In
evaluating
this
criterion,
EPA
looked
at
the
total
exposed
population,
as
well
as
the
population
exposed
above
the
estimated
HRL.
Estimates
of
the
populations
exposed
and
the
levels
to
which
they
are
exposed
were
derived
from
the
monitoring
results.
These
estimates
are
included
in
Chapter
4
of
this
document
and
are
summarized
in
Section
9.4.2.

In
order
to
evaluate
risk
from
exposure
through
drinking
water,
EPA
considered
the
net
environmental
exposure
in
comparison
to
the
exposure
through
drinking
water.
For
example,
if
exposure
to
a
contaminant
occurs
primarily
through
ambient
air,
regulation
of
emissions
to
air
provides
a
more
meaningful
opportunity
for
EPA
to
reduce
risk
than
regulation
of
the
contaminant
in
drinking
water.
In
making
the
preliminary
regulatory
determination,
the
available
information
on
exposure
through
drinking
water
(
Chapter
4)
and
information
on
exposure
through
other
media
(
Chapter
5)
were
used
to
estimate
the
fraction
that
drinking
water
contributes
to
the
total
exposure.
The
EPA
findings
are
discussed
in
Section
9.4.3.

In
making
its
preliminary
regulatory
determination,
EPA
also
evaluated
effects
on
potential
sensitive
populations,
including
the
fetus,
infants,
and
children.
The
sensitive
population
considerations
are
included
in
Section
9.4.4.

9.4.1
Risk
Criterion
Conclusions
The
data
discussed
in
this
section
and
Section
9.3.3
indicate
that
there
is
not
a
substantial
likelihood
that
aldrin
and
dieldrin
will
occur
in
public
water
systems
with
frequencies
and
at
levels
of
concern
for
public
health.

9.4.2
Exposed
Population
Estimates
As
noted
previously,
because
the
HRL
of
0.002
mg/
L
for
these
compounds
is
below
the
MRL,
any
recorded
detection
will
be
above
all
three
reference
levels
(
MRL,
HRL,
½
HRL).
Therefore,
estimates
of
the
national
population
exposed
to
concentrations
greater
than
any
of
these
levels
will
be
equivalent.
Summary
data
for
exposed
population
estimates
are
provided
below
in
Table
9­
3.
9­
18
Aldrin/
Dieldrin
 
February
2003
It
must
be
remembered
that
the
complete
R2­
ARS­
based
estimates
are
very
conservative
in
nature,
in
that
they
are
derived
from
a
collective
database
that
includes
incomplete
and
biased
state
data
sets,
and
because
only
a
single
detection
is
sufficient
to
classify
a
PWS
as
"
positive"
S
these
factors
will
tend
to
significantly
overestimate
the
true
sizes
of
the
exposed
populations.
On
the
other
hand,
using
data
only
from
the
Round
2
cross­
section
(
from
19
states,
the
R2­
X­
based
estimates),
which
have
been
screened
to
remove
incomplete,
biased,
and
otherwise
unusable
data
and
then
selected
to
geographically
represent
the
entire
nation,
is
less
likely
to
overestimate
and
may
even
underestimate
to
some
extent
the
potentially
exposed
national
populations.

For
aldrin,
the
median
and
99th
percentile
concentrations
of
detections
based
on
all
Round
2
UCM
data
were
0.18
and
4.40
:
g/
L,
respectively.
Based
only
on
the
19­
state
Round
2
crosssection
data,
the
corresponding
values
are
0.58
and
0.69
:
g/
L.
The
respective
two
sets
of
values
for
dieldrin
are
0.42
and
4.40
:
g/
L,
and
0.16
and
1.36
:
g/
L.
While
these
values
are
above
the
HRL
of
0.002
:
g/
L,
it
must
also
be
kept
in
mind
that
the
corresponding
values
for
all
samples
were
below
the
detection
limit,
and
that
the
HRL
itself
is
likely
a
very
conservative
estimate
of
any
human
risk
resulting
from
exposure
to
these
chemicals.

Table
9­
3.
National
Population
Estimates
for
Aldrin
and
Dieldrin
Exposure
via
Drinking
Water
Population
of
Concern
Round
2
Cross­
Section
(
19
States)
1
Round
2
Reporting
States2
Aldrin
Served
by
PWS
with
detections
38,871
1,051,989
Served
by
PWSs
with
detections
>
(
1/
2
HRL)
38,871
1,051,989
Served
by
PWSs
with
detections
>
HRL
38,871
1,051,989
Dieldrin
Served
by
PWS
with
detections
149,827
792,703
Served
by
PWSs
with
detections
>
(
1/
2
HRL)
149,827
792,703
Served
by
PWSs
with
detections
>
HRL
149,827
792,703
1Based
on
data
from
the
20­
State
Cross
Section,
minus
MA
(
SDWIS/
FED,
UCM
Round
2,
1993).
2Based
on
data
from
all
reporting
states
(
SDWIS/
FED,
UCM
Round
2,
1993).
Source:
Data
taken
from
Tables
4­
2
and
4­
5
in
Section
4.0
of
this
document.
Abbreviations:
HRL
=
Health
Reference
Level;
PWS
=
Public
Water
System.
9­
19
Aldrin/
Dieldrin
 
February
2003
9.4.3
Relative
Source
Contribution
Analysis
of
relative
source
contribution
compares
the
magnitude
of
exposures
(
i.
e.,
intakes)
expected
via
consumption
of
drinking
water
with
those
estimated
for
other
relevant
media
such
as
food,
air,
and
soil.
The
data
summarized
in
Chapter
4.0
provide
the
basis
for
estimating
the
amounts
of
aldrin
and
dieldrin
ingested
via
drinking
water
in
exposed
populations.
For
this
exercise,
the
non­
conservative
approach
was
taken
by
utilizing
the
median
and
99th
percentile
detect
concentrations
derived
from
only
UCM
Round
2
cross­
section
data
(
realizing
that
this
will
certainly
underestimate
to
some
degree
the
true
contribution
of
drinking
water
to
the
exposed
population's
total
intake
of
aldrin/
dieldrin).

For
a
70
kg
adult
consuming
2
L/
day
of
water
containing
aldrin
at
either
0.58
:
g/
L
(
median
detect
concentration)
or
0.69
:
g/
L
(
99th
percentile
detect
concentration),
the
corresponding
aldrin
intake
values
from
drinking
water
are
1.7
×
10­
5
and
2.0
×
10­
5
mg/
kg
bw/
day,
respectively.
For
a
10
kg
child
consuming
1
L/
day
of
water,
the
comparable
values
are
5.8
×
10­
5
and
6.9
×
10­
5
mg/
kg
bw/
day.

Similarly,
for
median
and
99th
percentile
detect
concentrations
of
dieldrin
(
0.16
and
1.36
:
g/
L,
respectively),
the
corresponding
adult
drinking
water
intake
values
of
dieldrin
are
0.46
×
10­
5
and
3.9
×
10­
5
mg/
kg
bw/
day,
respectively.
Dieldrin
drinking
water
intake
values
for
the
10
kg
child
are
1.6
×
10­
5
and
14
×
10­
5
mg/
kg
bw/
day.

Chapter
5
presents
data
on
the
estimated
daily
dietary
intake
of
aldrin
and
dieldrin
(
see
Tables
5­
3
and
5­
4).
Combining
estimates
for
non­
fish
food
with
those
for
fish
and
shellfish,
adult
and
child
dietary
intakes
of
aldrin
are
estimated
at
3.3
to
6.5
×
10­
5
and
13
to
18
×
10­
5
mg/
kg
bw/
day,
respectively.
For
dieldrin,
the
comparable
adult
and
child
dietary
intakes
are
3.6
×
10­
5
and
14
×
10­
5
mg/
kg
bw/
day.

Comparing
these
derived
estimates
for
intakes
via
drinking
water
and
diet,
the
ratios
of
dietary
intake
to
drinking
water
intake
for
aldrin
range
from
1.7
to
3.8
across
all
combinations
of
age
and
drinking
water
concentration
level.
For
dieldrin,
the
food/
water
intake
ratios
for
adults
and
children
are
0.9
and
1.0
using
the
99th
percentile
water
concentration,
and
7.8
and
8.8
using
the
median
water
concentration.
Applying
the
more
"
conservative"
aldrin/
dieldrin
water
concentrations
based
on
the
monitoring
data
of
all
reporting
UCM
Round
2
states
would
reduce
all
of
these
food/
water
ratios
by
a
factor
of
approximately
3
to
6.
Thus,
when
conservatively
analyzed
relative
to
the
diet,
drinking
water
could
potentially
be
responsible
for
a
significant
portion
of
total
daily
intake
of
aldrin/
dieldrin,
but
only
for
limited
populations
under
exposure
circumstances
that
are
considered
unlikely.

Referring
again
to
Tables
5­
3
and
5­
4,
it
can
be
seen
that
the
estimated
daily
intakes
of
aldrin
and
dieldrin
from
air
for
adults
and
children
range
from
0.013
×
10­
5
to
0.24
×
10­
5
mg/
kg
bw/
day.
Despite
the
fact
that
these
values
are
likely
significant
overestimates
since
they
are
based
on
data
that
is
30
years
old,
they
are
still
small
relative
to
drinking
water
and
dietary
intakes.
Although
soil
data
were
not
available
for
aldrin,
those
for
dieldrin
indicate
that
ingestion
of
soil
represents
only
a
minor
exposure
pathway
for
these
compounds.
9­
20
Aldrin/
Dieldrin
 
February
2003
9.4.4
Sensitive
Populations
The
issue
of
sensitive
populations
has
already
been
addressed
to
the
extent
currently
possible.
While
there
is
some
reasonable
basis
to
suspect
that
fetuses,
young
children,
the
elderly,
and
those
having
compromised
liver,
immune,
or
even
neurological
function
may
be
at
increased
risk
for
one
or
more
of
the
toxic
effects
of
aldrin/
dieldrin,
such
susceptibility
has
not
yet
been
convincingly
demonstrated
or
adequately
quantified
in
the
scientific
literature.

9.5
Regulatory
Determination
Summary
While
there
is
evidence
that
aldrin/
dieldrin
may
have
adverse
health
effects,
including
the
probability
to
cause
cancer
in
humans,.
neither
contaminant
has
been
used
in
the
US
since
1987.
Furthermore,
monitoring
data
indicate
that
the
contaminants'
concentrations
have
been
declining
since
the
cancellation
of
their
registrations
as
pesticides.
Their
occurrences
in
public
water
systems
have
also
been
very
limited
and
at
very
low
concentrations.
For
these
reasons,
regulation
of
aldrin
and
dieldrin
may
not
present
a
meaningful
opportunity
for
health
risk
reduction
for
persons
served
by
public
water
systems.
Therefore,
EPA
may
not
propose
to
regulate
aldrin/
dieldrin
with
NPDWRs.
All
final
determinations
and
future
analysis
will
be
presented
in
the
Federal
Register
Notice
covering
CCL
proposals.
9­
21
Aldrin/
Dieldrin
 
February
2003
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Korte.
1974.
Fate
of
aldrin­
14C
in
maize,
wheat
and
soils
under
outdoor
conditions.
J.
Agric.
Food
Chem.
22(
4):
609­
612
(
as
cited
in
HSDB,
2000a;
IPCS,
1989b).

Winteringham,
F.
P.
W.
and
J.
B.
Barnes.
1955.
Comparative
response
of
insects
and
mammals
to
certain
halogenated
hydrocarbons
used
as
pesticides.
Physiol.
Rev.
35:
701
(
as
cited
in
USEPA,
1980).

Witherup,
S.,
K.
L.
Stemmer,
J.
L.
Roberts,
et
al.
1961.
Prolonged
cutaneous
contact
of
wool
impregnated
with
dieldrin.
The
Kettering
Laboratory
in
the
Department
of
Preventive
Medicine
and
Industrial
Health,
College
of
Medicine,
University
of
Cincinnati.
Cincinnati,
OH
(
as
cited
in
ATSDR,
2000).

Wolff,
T.,
E.
Deml
and
H.
Wanders.
1979.
Aldrin
epoxidation,
a
highly
sensitive
indicator
specific
for
cytochrome
P­
450­
dependent
monooxygenase
activities.
Drug
Metab.
Dispos.
7:
301­
305
(
as
cited
in
ATSDR,
2000).

Wong,
D.
T.
and
L.
C.
Terriere.
1965.
Epoxidation
of
aldrin,
isodrin,
and
heptachlor
by
rat
liver
microsomes.
Biochem.
Pharmacol.
14:
375­
377
(
as
cited
in
ATSDR,
2000;
IARC,
1974a;
USEPA,
1980).

Woolley,
D.,
L.
Zimmer,
D.
Dodge,
and
K.
Swanson.
1985.
Effects
of
lindane­
type
insecticides
in
mammals:
Unsolved
problems.
Neurotoxicity
6:
165­
192
(
as
cited
in
ATSDR,
2000).

Wright,
A.
S.,
C.
Donninger,
R.
D.
Greenland,
K.
L.
Stemmer
and
M.
R.
Zavon.
1978.
The
effects
of
prolonged
ingestion
of
dieldrin
on
the
livers
of
male
rhesus
monkeys.
Ecotoxicol.
Environ.
Saf.
1(
4):
477­
502
(
as
cited
in
IPCS,
1989b;
TOXLINE,
2000a).

Wright,
A.
S.,
D.
Potter,
M.
F.
Wooder,
C.
Donninger,
R.
D.
Greenland.
1972.
The
effects
of
dieldrin
on
the
subcellular
structure
and
function
of
mammalian
liver
cells.
Food
Cosmet.
Toxicol.
10:
311­
332
(
as
cited
in
ATSDR,
2000).

Yess,
N.
J.,
E.
L.
Gunderson,
and
R.
R.
Roy.
1993.
U.
S.
Food
and
Drug
Administration
monitoring
of
pesticide
residues
in
infant
foods
and
adult
foods
eaten
by
infants/
children.
J.
AOAC
Int.
76
(
3):
492­
507.
10­
30
Aldrin/
Dieldrin
 
February
2003
Zabik,
M.
E.,
M.
J.
Zabik,
A.
M.
Booren,
S.
Daubenmire,
M.
A.
Pascall,
R.
Welch,
and
H.
Humphrey.
1995.
Pesticides
and
total
polychlorinated
biphenyls
residues
in
raw
and
cooked
walleye
and
white
bass
harvested
from
the
Great
Lakes.
Bull.
Environ.
Contam.
Toxicol.
54
(
3):
396­
402.

Zhong­
Xiang,
L.,
T.
Kavanagh,
J.
E.
Trosko,
and
C.
C.
Chang.
1986.
Inhibition
of
gap
junctional
intercellular
communication
in
human
teratocarcinoma
cells
by
organochlorine
pesticides.
Toxicol.
Appl.
Pharmacol.
83:
10­
19
(
as
cited
in
GAP2000,
2000a,
b).
A1
Aldrin/
Dieldrin
 
February
2003
Abbreviations
and
Acronyms
ACGIH
­
American
Conference
of
Governmental
Industrial
Hygienists
ADI
­
Acceptable
Daily
Intake
AI
­
active
ingredient
ATSDR
­
Agency
for
Toxic
Substances
and
Disease
Registry
BCF
­
bioconcentration
factor
BCH
­
bicycloheptadiene
BTFs
­
biotransfer
factors
CASRN
­
Chemical
Abstract
Service
Registry
Number
CCL
­
Contaminant
Candidate
List
CERCLA
­
Comprehensive
Environmental
Response,
Compensation
&
Liability
Act
CI
­
Confidence
Interval
CMR
­
Chemical
Monitoring
Reform
CNS
­
central
nervous
system
CWSs
­
community
water
systems
DBCP
­
dibromochloropropane
2,3­
DHBA
­
2,3­
dihydroxybenzoic
acid
DMPC
­
dimyristoylphosphatidylcholine
DNA
­
deoxyribonucleic
acid
DPH
­
1,6­
diphenyl­
1,3,5­
hexatriene
DPH­
PA
­
propionic
acid
deriative
of
DPH
DRG
­
dorsal
root
ganglion
EEG
­
electroencephalogram
EMAP
­
Environmental
Monitoring
Assessment
Program
EPA
­
Environmental
Protection
Agency
EPCRA
­
Emergency
Planning
and
Community
Right­
to­
Know
Act
F
­
female
FDA
­
Food
and
Drug
Administration
FIFRA
­
Federal
Insecticide,
Fungicide,
and
Rodenticide
Act
FSH
­
follicle
stimulating
hormone
GABA
­
gamma
aminobutyric
acid
GAD­
ir
­
glutamate
decarboxylase
immunoreactive
GC/
MS
­
gas
chromotography/
mass
spectometry
gd
­
gestation
day
GI
­
gastrointestinal
tract
G6P­
­
glucose­
6­
phosphatase
deficient
GW
­
ground
water
A2
Aldrin/
Dieldrin
 
February
2003
HazDat
­
Hazardous
Substance
Release
and
Health
Effects
Database
HCCPD
­
hexachlorocyclopentadiene
HRL
­
Health
Reference
Level
HSDB
­
Hazardous
Substances
Data
Bank
IARC
­
International
Agency
for
Research
on
Cancer
IC
50
­
IOC
­
inorganic
contaminant
IPCS
­
International
Programme
on
Chemical
Safety
IRIS
­
Integrated
Risk
Information
System
IUPAC
­
International
Union
of
Pure
and
Applied
Chemistry
LD
50
­
lethal
dose
LH
­
Lutenizing
hormone
LOAEL
­
lowest­
observed­
adverse­
effect
level
M
­
male
MCLG
­
Maximum
Contaminant
Level
Goal
MDA
­
malondialdehyde
MDL
­
Method
Detection
Limit
MMT
­
methylcyclopentadienyl
manganese
tricarbonyl
MRL
­
Minimum
Reporting
Level
mRNA
­
messenger
ribonucleic
acid
MW
­
molecular
weight
NADPH
­
nicotine
adenine
dinucleotide
phosphate
NAS/
OW
­
National
Academy
of
Sciences/
Office
of
Water
NAWQA
­
National
Water
Quality
Assessment
Program
NCOD
­
National
Drinking
Water
Contaminant
Occurrence
Database
NDWAC
­
National
Drinking
Water
Advisory
Council
NOAA
­
National
Oceanic
and
Atmospheric
Administration
NOAEL
­
no­
observed­
adverse­
effect
level
NPDWR
­
National
Primary
Drinking
Water
Regulation
NPL
­
National
Priorities
List
NTNCWSs
­
non­
purchased
non­
transient
non­
community
water
systems
OH8dG
­
8­
hydroxy­
2'­
deoxyguanosine
PB
­
phenobarbital
PES
­
prostaglandin
endoperoxide
synthase
PGG
2
­
prostaglandin
G
2
PHG
2
­
prostaglandin
H
2
ppd
­
postpartum
day
ppm
­
part
per
million
A3
Aldrin/
Dieldrin
 
February
2003
PWS
­
Public
Water
System
q1*
­
geometric
mean
R2­
ARS
­
Round
2
states
RBCs
­
red
blood
cells
RCRA
­
Resource
Conservation
and
Recovery
Act
RfD
­
Reference
Dose
ROS
­
reactive
oxygen
species
RSD
­
risk­
specific
dose
R2­
X
­
Round
2
cross­
section
SARA
Title
III
­
Superfund
Amendments
and
Reauthorization
Act
SCE
­
sister
chromatid
exchanges
SDWA
­
Safe
Drinking
Water
Act
SDWIS/
FED
­
Safe
Drinking
Water
Information
System
(
Federal
version)
SMRs
­
standardized
mortality
ratios
SOC
­
synthetic
organic
compound
35S­
TBPS
­
t­
35S
butyl­
bicyclophosphorothionate
SW
­
surface
water
TE
50
­
median
effective
time
TH­
ir
­
tyrosine
hydroxylase­
immunoreactive
TRI
­
Toxic
Release
Inventory
UCM
­
Unregulated
Contaminant
Monitoring
UCMR
­
Unregulated
Contaminant
Monitoring
Regulation/
Rule
UDPGA
­
uridine
diphosphoglucuronic
acid
UDS
­
unscheduled
DNA
synthesis
URCIS
­
Unregulated
Contaminant
Monitoring
Information
System
USDA
­
United
States
Department
of
Agriculture
USEPA
­
United
States
Environmental
Protection
Agency
USGS
­
United
States
Geological
Survey
UV
­
ultraviolet
VOC
­
volatile
organic
compound
atm
­
atmospheres
atm­
m3/
mol
­
atmospheres
cubic
meter
per
mole
oC
­
degrees
Celsius
cm
­
centimeters
cm2
­
square
centimeters
g
­
grams
g/
cc
­
grams
per
cubic
centimeter
kg
­
kilograms
A4
Aldrin/
Dieldrin
 
February
2003
kg/
day
­
kilograms
per
day
kg/
ha
­
kilograms
per
hectare
L/
day
­
liters
per
day
lbs
­
pounds
M
­
molar
m
­
meter
mg/
cm2
­
milligrams
per
square
centimeter
mg/
day
­
milligrams
per
day
mg/
kg
­
milligrams
per
kilogram
mg/
kg
bw
­
milligrams
per
kilogram
per
body
weight
mg/
kg
bw/
day
­
milligrams
per
kilogram
per
body
weight
per
day
mg/
kg
bw/
week
­
miiligrams
per
kilogram
per
body
weight
per
week
mg/
L
­
milligrams
per
liter
mg/
m3
­
milligrams
per
cubic
meter
mL
­
milliliter
mm
Hg
­
millimeters
of
mercury
ng/
g
­
nanograms
per
gram
ng/
L
­
nanograms
per
liter
ng/
m3
­
nanograms
per
cubic
meter
ng/
mL
­
nanograms
per
milliliter
nM
­
nanomolar
nm
­
nanometers
nmol/
mL
­
nanomole
per
milliliter
pM
­
pico
molar
ppb
­
parts
per
billion
:
g
­
micrograms
:
g/
cm2
­
micrograms
per
square
centimeter
:
g/
g
­
micrograms
per
gram
:
g/
L
­
micrograms
per
liter
:
g/
m2
­
micrograms
per
square
meter
:
g/
m3
­
micrograms
per
cubic
meter
:
M
­
micro
molar
B1
Aldrin/
Dieldrin
 
February
2003
APPENDIX
A:
Round
2
Aldrin
Occurrence
Aldrin
Occurrence
in
Public
Water
Systems
in
Round
2,
UCM
(
1993)
results
STATE
TOTAL
UNIQUE
PWS
#
GW
PWS
#
SW
PWS
%
PWS
with
detections
%
GW
PWS
with
%
SW
PWS
with
detections
%
PWS
>
HRL
%
GW
PWS
>
HRL
%
SW
PWS
>
HRL
99%
VALUE
(
µ
g/
L)

Tribes
(
06)
26
25
1
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.50
AK
34
24
10
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
AL
16
11
5
100.00%
100.00%
100.00%
100.00%
100.00%
100.00%
0.68
AR
536
431
105
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
AZ
CA
CO
750
538
212
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
CT
70
35
35
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
IN
KY
366
184
182
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
2.00
LA
1,363
1,295
68
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.01
MA
56
29
27
17.86%
17.24%
18.52%
17.86%
17.24%
18.52%
4.40
MD
726
669
57
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
1.00
ME
MI
2,650
2,570
80
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
MN
1,264
1,234
30
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
MO
378
280
98
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.10
MS
12
11
1
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
NC
536
490
46
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
ND
296
258
38
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.01
NH
593
560
33
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
NJ
NM
720
691
29
0.14%
0.14%
0.00%
0.14%
0.14%
0.00%
<
1.00
OH
1,029
882
147
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
30.00
OK
98
76
22
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
OR
1,152
999
153
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
PA
68
57
11
5.88%
7.02%
0.00%
5.88%
7.02%
0.00%
0.10
RI
24
15
9
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.20
SC
939
841
98
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
SD
TN
7
2
5
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
TX
427
122
305
0.23%
0.82%
0.00%
0.23%
0.82%
0.00%
<
0.20
VT
401
349
52
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
WA
586
517
69
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
WI
TOTAL
15,123
13,195
1,928
0.21%
0.17%
0.52%
0.21%
0.17%
0.52%
<
1.00
20
STATES
12,221
10,569
1,652
0.10%
0.07%
0.30%
0.10%
0.07%
0.30%
<
2.00
19
STATES1
12,165
10,540
1,625
0.02%
0.02%
0.00%
0.02%
0.02%
0.00%
<
2.00
1.
Massachusetts
data
not
included
in
"
19
States"
summary
statistics
for
Aldrin.

PWS=
Public
Water
Systems;
GW=
Ground
Water
(
PWS
Source
Water
Type);
SW=
Surface
Water
(
PWS
Source
Water
Type);
MRL=
Minimum
Reporting
Limit
(
for
laboratory
analyses).

The
Health
Reference
Level
(
HRL)
is
the
estimated
health
effect
level
as
provided
by
EPA
for
preliminary
assessment
for
this
work
assignment.

"%
>
HRL"
indicates
the
proportion
of
systems
with
any
analytical
results
exceeding
the
concentration
value
of
the
HRL.

The
Health
Reference
Level
(
HRL)
used
for
Aldrin
is
0.002
µ
g/
L.
This
is
a
draft
value
for
working
review
only.

The
highlighted
States
are
part
of
the
SDWIS/
FED
20
State
Cross­
Section.
B2
Aldrin/
Dieldrin
 
February
2003
APPENDIX
B:
Round
2
Dieldrin
Occurrence
Dieldrin
Occurrence
in
Public
Water
Systems
in
Round
2,
UCM
(
1993)
results
STATE
TOTAL
UNIQUE
PWS
#
GW
PWS
#
SW
PWS
%
PWS
>
MRL
%
GW
PWS
>
MRL
%
SW
PWS
>
MRL
%
PWS
>
HRL
%
GW
PWS
>
HRL
%
SW
PWS
>
HRL
99%
VALUE
(
µ
g/
L)

Tribes
(
06)
25
24
1
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.10
AK
16
12
4
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
AL
4
4
0
100.00%
0.00%
0.00%
100.00%
100.00%
0.00%
0.10
AR
536
431
105
0.19%
0.00%
0.95%
0.19%
0.00%
0.95%
<
0.00
AZ
CA
CO
749
537
212
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
CT
70
35
35
1.43%
0.00%
2.86%
1.43%
0.00%
2.86%
<
0.00
IN
KY
44
20
24
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.21
LA
1,363
1,295
68
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.07
MA
55
28
27
18.18%
17.86%
18.52%
18.18%
17.86%
18.52%
4.40
MD
725
668
57
0.97%
0.90%
1.75%
0.97%
0.90%
1.75%
<
1.00
ME
MI
2,650
2,570
80
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
MN
1,264
1,234
30
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
MO
378
280
98
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.10
MS
12
11
1
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
NC
522
475
47
0.38%
0.42%
0.00%
0.38%
0.42%
0.00%
<
0.00
ND
296
258
38
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.01
NH
593
560
33
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
NJ
NM
716
687
29
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.20
OH
1,029
883
146
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
20.00
OK
98
76
22
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
OR
1,148
995
153
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
PA
67
56
11
7.46%
8.93%
0.00%
7.46%
8.93%
0.00%
0.10
RI
15
6
9
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.30
SC
939
841
98
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
SD
TN
7
2
5
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
TX
427
122
305
0.23%
0.82%
0.00%
0.23%
0.82%
0.00%
<
0.20
VT
395
343
52
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
WA
582
515
67
0.00%
0.00%
0.00%
0.00%
0.00%
0.00%
<
0.00
WI
TOTAL
14,725
12,968
1,757
0.21%
0.18%
0.46%
0.21%
0.18%
0.46%
<
0.30
20
STATES
11,843
10,357
1,486
0.18%
0.14%
0.47%
0.18%
0.14%
0.47%
<
1.00
19
STATES
11,788
10,329
1,459
0.09%
0.09%
0.14%
0.09%
0.09%
0.14%
<
1.00
1.
Massachusetts
data
not
included
in
"
19
States"
summary
statistics
for
Dieldrin.

PWS=
Public
Water
Systems;
GW=
Ground
Water
(
PWS
Source
Water
Type);
SW=
Surface
Water
(
PWS
Source
Water
Type);
MRL=
Minimum
Reporting
Limit
(
for
laboratory
analyses).

The
Health
Reference
Level
(
HRL)
is
the
estimated
health
effect
level
as
provided
by
EPA
for
preliminary
assessment
for
this
work
assignment.

"%
>
HRL"
indicates
the
proportion
of
systems
with
any
analytical
results
exceeding
the
concentration
value
of
the
HRL.

The
Health
Reference
Level
(
HRL)
used
for
Dieldrin
is
0.002
µ
g/
L.
This
is
a
draft
value
for
working
review
only.
The
highlighted
States
are
part
of
the
SDWIS/
FED
20
State
Cross­
Section.