United States
Environmental Protection
Agency
Health Effects Support
Document for 1,1-Dichloro-2,2
bis(p-chlorophenyl)ethylene
(DDE)
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Health Effects Support Document
for
l,l-Dichloro-2,2-bis(/7-chlorophenyl)ethylene (DDE)
U.S. Environmental Protection Agency
Office of Water (43 04T)
Health and Ecological Criteria Division
Washington, DC 20460
www.epa.gov/safewater/ccl/pdf/DDE.pdf
EPA Document Number EPA-822-R-08-003
January, 2008
Printed on Recycled Paper
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DDE — January, 2008 IV
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FOREWORD
The Safe Drinking Water Act (SDWA), as amended in 1996, requires the Administrator
of the Environmental Protection Agency (EPA) to establish a list of contaminants to aid the
Agency in regulatory priority setting for the drinking water program. In addition, the SDWA
requires EPA to make regulatory determinations for no fewer than five contaminants by August
2001 and every five years thereafter. The criteria used to determine whether or not to regulate a
chemical on the Contaminant Candidate List (CCL) are the following:
• 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 criteria are used in making a determination to
regulate a contaminant. The Agency may determine that there is no need for regulation when a
contaminant fails to meet one of the criteria. The 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 regulatory determination for
DDE. In arriving at the regulatory determination, 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 (Agency for Toxic Substances and Disease Registry). 2002. Toxicological
profile for DDT, DDE, and ODD (2002 update). Department of Health and Human
Services. Available from: . MRLs
posted at: .
U.S. EPA (United States Environmental Protection Agency). 1988b. Integrated Risk
Information System (IRIS):/\p'-Dichlorodiphenyldichloroethylene (DDE). Cincinnati,
OH. Available from: .
WHO (World Health Organization). 2004. DDT and its derivatives in drinking water:
Background document for the development of the WHO guideline for drinking water
quality. Available from:
.
DDE —January, 2008
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Information from the published risk assessments was supplemented with information
from the primary references for key studies and recent studies of DDE identified by a literature
search conducted in 2004.
A reference dose (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, significant body or organ weight changes, blood disorders, etc. 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 oral 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 DDE includes a formal hazard identification and an
estimate of tumorigenic potency when available. Hazard identification is a weight-of-evidence
judgment of the likelihood that the agent is a human carcinogen via the oral route and of the
conditions under which the carcinogenic effects may be expressed.
Development of these hazard identification and dose-response assessments for DDE has
followed the general guidelines for risk assessment as set forth by the National Research Council
(1983). EPA guidelines that were used in the development of this assessment may include the
following: Guidelines for the Health Risk Assessment of Chemical Mixtures (U.S. EPA, 1986a),
Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986b), Guidelines for Developmental
Toxicity Risk Assessment (U.S. EPA, 1991), Guidelines for Reproductive Toxicity Risk
Assessment (U.S. EPA, 1996), Guidelines for Neurotoxicity Risk Assessment (U.S. EPA, 1998a),
Guidelines for Carcinogen Assessment (U.S. EPA 2005a), Recommendations for and
Documentation of Biological Values for Use in Risk Assessment (U.S. EPA, 1988a), (proposed)
Interim Policy for Particle Size and Limit Concentration Issues in Inhalation Toxicity (U.S.
EPA, 1994a), Methods for Derivation of Inhalation Reference Concentrations and Application of
Inhalation Dosimetry (U.S. EPA, 1994b), Use of the Benchmark Dose Approach in Health Risk
Assessment (U.S. EPA, 1995), Science Policy Council Handbook: Peer Review (U.S. EPA,
1998b, 2000a), Science Policy Council Handbook: Risk Characterization (U.S. EPA, 2000b),
Benchmark Dose Technical Guidance Document (U.S. EPA, 2000c), Supplementary Guidance
for Conducting Health Risk Assessment of Chemical Mixtures (U.S. EPA, 2000d), and^4 Review
of the Reference Dose and Reference Concentration Processes (U.S. EPA, 2002a).
The chapter on occurrence and exposure to DDE through potable water was developed by
the Office of Ground Water and Drinking Water. It is based primarily on first Unregulated
Contaminant Monitoring Rule (UCMR 1) data collected under the SDWA. The UCMR 1 data
are supplemented with ambient water data, as well as data from the States, and published papers
on occurrence in drinking water.
DDE — January, 2008 VI
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ACKNOWLEDGMENT
This document was prepared under the U.S. EPA contract No. 68-C-02-009, Work
Assignment No. 2-54, 3-54, and 4-54 with ICF International, Fairfax, VA. The Lead U.S. EPA
Scientist was Brenda Foos, Health and Ecological Criteria Division, Office of Science and
Technology, Office of Water.
DDE — January, 2008 Vll
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DDE — January, 2008 Vlll
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TABLE OF CONTENTS
FOREWORD v
ACKNOWLEDGMENT vii
LIST OF TABLES xiii
LIST OF FIGURES xv
1.0 EXECUTIVE SUMMARY 1-1
2.0 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES 2-1
3.0 USES AND ENVIRONMENTAL FATE 3-1
3.1 Production and Use 3-1
3.2 Environmental Release 3-1
3.3 Environmental Fate 3-2
3.4 Summary 3-5
4.0 EXPOSURE FROM DRINKING WATER 4-1
4.1 Introduction 4-1
4.2 Ambient Occurrence 4-1
4.2.1 Data Sources and Methods 4-1
4.2.2 Results 4-3
4.3 Drinking Water Occurrence 4-7
4.3.1 Data Sources and Methods 4-7
4.3.2 Derivation of the Health Reference Level 4-8
4.3.3 Results 4-8
4.4 Summary 4-12
5.0 EXPOSURE FROM MEDIA OTHER THAN WATER 5-1
5.1 Exposure from Food 5-1
5.1.1 Concentration in Non-Fish Food Items 5-1
5.1.2 Concentrations in Fish, Shellfish and Marine Mammals 5-3
5.1.3 Intake of DDE from Food 5-5
5.2 Exposure from Air 5-6
5.2.1 Concentration of DDE in Air 5-6
5.2.2 Intake of DDE from Air 5-8
5.3 Exposure from Soil 5-8
5.3.1 Concentration of DDE in Soil 5-9
5.3.2 Intake of DDE from Soil 5-11
5.4 Other Residential Exposures 5-11
5.5 Occupational (Workplace) Exposures 5-12
DDE — January, 2008 ix
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5.6 Summary 5-12
6.0 TOXICOKINETICS 6-1
6.1 Absorption 6-1
6.2 Distribution 6-1
6.3 Metabolism 6-3
6.4 Excretion 6-4
7.0 HAZARD IDENTIFICATION 7-1
7.1 Human Effects 7-1
7.1.1 Short-Term Studies and Case Reports 7-1
7.1.2 Long-Term and Epidemiological Studies 7-1
7.2 Animal Studies 7-10
7.2.1 Acute Toxicity 7-10
7.2.2 Short-Term Studies 7-10
7.2.3 Subchronic Studies 7-11
7.2.4 Neurotoxicity 7-12
7.2.5 Developmental/Reproductive Toxicity 7-12
7.2.6 Chronic Toxicity 7-14
7.2.7 Carcinogenicity 7-16
7.3 Other Key Data 7-18
7.3.1 Mutagenicity and Genotoxicity 7-18
7.3.2 Immunotoxicity 7-18
7.3.3 Hormonal Disruption 7-19
7.3.4 Physiological or Mechanistic Studies 7-20
7.3.5 Structure-Activity Relationship 7-22
7.4 Hazard Characterization 7-22
7.4.1 Synthesis and Evaluation of Major Noncancer Effects 7-22
7.4.2 Synthesis and Evaluation of Carcinogenic Effects 7-22
7.4.3 Mode of Action and Implications in Cancer Assessment 7-23
7.4.4 Weight of Evidence Evaluation for Carcinogenicity 7-23
7.4.5 Potentially Sensitive Populations 7-23
8.0 DOSE-RESPONSE ASSESSMENT 8-1
8.1 Dose-Response for Noncancer Effects 8-1
8.2 Dose-Response for Cancer Effects 8-1
8.2.1 Choice of Study 8-1
8.2.2 Dose-Response Characterization 8-3
8.2.3 Extrapolation Models and Rationale 8-3
8.2.4 Cancer Potency and Unit Risk 8-4
9.0 REGULATORY DETERMINATION AND CHARACTERIZATION OF RISK FROM
DRINKING WATER 9-1
9.1 Regulatory Determination for Chemicals on the CCL 9-1
9.1.1 Criteria for Regulatory Determination 9-1
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9.1.2 National Drinking Water Advisory Council Recommendations 9-2
9.2 Health Effects 9-2
9.2.1 Health Criterion Conclusion 9-3
9.2.2 Hazard Characterization and Mode of Action Implications 9-3
9.2.3 Dose-Response Characterization and Implications in Risk Assessment
9-4
9.3 Occurrence in Public Water Systems 9-4
9.3.1 Occurrence Criterion Conclusion 9-4
9.3.2 Monitoring Data 9-5
9.3.3 Use and Fate Data 9-6
9.4 Risk Reduction 9-6
9.4.1 Risk Criterion Conclusion 9-7
9.4.2 Exposed Population Estimates 9-7
9.4.3 Relative Source Contribution 9-8
9.4.4 Sensitive Populations 9-8
9.5 Regulatory Determination Decision 9-8
10.0 REFERENCES 10-1
APPENDIX A: Abbreviations and Acronyms Appendix A-l
APPENDIX B: HMD Modeling Output Appendix B-l
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DDE — January, 2008 Xll
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LIST OF TABLES
Table 2-1 Chemical and Physical Properties of/?,/>'-Dichlorodiphenyldichloroethylene . 2-2
Table 4-1 USGS Pesticide National Synthesis Summary of NAWQA Monitoring ofp,p'-
DDE in Ambient Surface Water, 1992-2001 4-3
Table 4-2 USGS Pesticide National Synthesis Summary of NAWQA Monitoring ofp,p'-
DDE in Ambient Ground Water, 1992-2001 4-4
Table 4-3 USGS Pesticide National Synthesis Summary of NAWQA Monitoring ofp,p'-
DDE in Bed Sediment, 1992-2001 4-5
Table 4-4 USGS Pesticide National Synthesis Summary of NAWQA Monitoring ofp,p'-
DDE in Whole Fish, 1992-2001 4-5
Table 4-5 USGS Pesticide National Synthesis Summary of NAWQA Monitoring of o,p'-
DDE in Bed Sediment, 1992-2001 4-6
Table 4-6 USGS Pesticide National Synthesis Summary of NAWQA Monitoring ofo,p'-
DDE in Whole Fish, 1992-2001 4-7
Table 4-7 Summary UCMR 1 Occurrence Statistics for 4,4'-DDE in Small Systems (Based
on Statistically Representative National Sample of Small Systems) 4-10
Table 4-8 Summary UCMR 1 Occurrence Statistics for 4,4'-DDE in Large Systems (Based
on the Census of Large Systems) 4-11
Table 5-1 Percent of food samples with detectable DDE levels 5-2
Table 5-2 Concentration of DDE in fish, shellfish, and marine mammals 5-4
Table 5-3 Mean Daily Intake of DDE from Food Per Unit Body Weight (|ig/kg body
weight/day) for Various Age Groups in the United States 5-5
Table 5-4 DDE Concentrations in North Dakota Air Sampling Collected in 1993 and 1994
5-7
Table 7-1 Tumors observed in the Rossi et al. (1983) study in Syrian Golden Hamsters
7-17
Table 7-2 Summary of NTP (2005) Genetic Toxicology Results 7-18
Table 8-1 Dose Response Data for the Carcinogenicity of DDE 8-2
Table 8-2 Summary of Liver Tumor Incidence, 78-Week Study in Mice 8-3
Table 8-3 Factors Used to Derive the Previous Oral Slope Factor for DDE 8-5
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DDE — January, 2008 XIV
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LIST OF FIGURES
Figure 2-1 Chemical Structure ofp,//-Dichlorodiphenyldichloroethylene 2-1
Figure 4-1 Geographic Distribution of 4,4'-DDE in UCMR 1 Monitoring — States With at
Least One Detection at or Above the MRL (0.8 |ig/L) 4-12
Figure 6-1 Proposed metabolic pathway for DDT, including further metabolism of DDE
(adapted from Peterson and Robinson, 1964 and presented in ATSDR, 2002)
6-5
Figure 6-2 Proposed metabolic pathway for the conversion ofp,p'-DDE to its methylsulfone
derivatives (Bergman et al., 1994; Letcher et al., 1998; Weistrand and Noren,
1997) 6-6
Figure 8-1 Multistage (2) Model with 0.95 Confidence Interval for the Female Mice Data
from the NCI (1978) Tumor Bioassay 8-4
DDE — January, 2008 XV
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DDE — January, 2008 XVI
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1.0 EXECUTIVE SUMMARY
The U.S. Environmental Protection Agency (EPA) has prepared this Health Effects
Support Document for l,l-Dichloro-2,2-bis(p-chlorophenyl)ethylene (DDE) to assist in
determining whether to regulate DDE with a National Primary Drinking Water Regulation
(NPDWR). The available data on occurrence, exposure, and other risk considerations suggest
that, because DDE does not occur in public water systems at frequencies and levels of public
health concern, regulating DDE will not present a meaningful opportunity to reduce health risk.
EPA will present a determination and further analysis in the Federal Register Notice covering the
Contaminant Candidate List (CCL) regulatory determinations.
DDE (Chemical Abstracts Services Registry Number 72-55-9) has never been produced
commercially, but is a primary environmental and metabolic degradation product of the pesticide
DDT. Although all uses of DDT in the U.S. were cancelled on January 1, 1973, DDT and DDE
remain in the environment because they are persistent and bioaccumulative. Also, DDT is still
used in other parts of the world.
Occurrence data indicate that DDE was not detected at concentrations above the
minimum reporting level (MRL; 0.8 |ig/L) in any small public water systems. DDE was
detected at one monitored large groundwater system. This system represented 0.03% of large
public water systems and 0.01% of the population served by them (approximately 18,000
people). Because the MRL for DDE monitoring is greater than half the Health Reference Level
(/^HRL or 0.1 |ig/L) and the full HRL (HRL or 0.2 |ig/L), it cannot be stated how many samples
may have contained DDE at the HRL or /^ the HRL. DDE has been detected in ambient surface
water samples, but none of the detections were greater than /^HRL, or the full HRL.
Accordingly, DDE is not likely to occur in public water systems at concentrations of concern.
Moreover, based on the occurrence of DDE in food items (at concentrations up to 0.102 ppm)
and the estimated DDE intake due to food consumption (up to 0.0441 ng/kg body weight/day),
foods are likely a greater source of DDE exposure than drinking water.
Health effects information has identified both cancer and non-cancer effects associated
with exposure to DDE. DDE has been associated with reduced body weight gain, and
neurological, liver and kidney effects in laboratory animals. DDE has also been found to be an
antiandrogenic compound, which may explain a number of reproductive and developmental
effects seen in male rats exposed to DDE at various ages. Oral exposure to DDE has been
associated with increases in the incidence of liver tumors, including carcinomas, in two strains of
mice and in hamsters, as well as nonsignificant increases in the incidence of thyroid tumors in
female rats and adrenal tumors in hamsters. There have also been several human epidemiology
studies of DDE and cancer, but overall these studies have yielded conflicting results.
The DDE HRL of 0.2 |ig/L is based on female mouse hepatocellular carcinomas data.
This number was derived using the oral slope factor of 1.7 x 10"1 (mg/kg-day)"1 and default
exposure assumptions of 70 kg body weight and 2 L per day drinking water ingestion.
DDE — January, 2008 1-1
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2.0 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES
l,l-Dichloro-2,2-bis(p-chlorophenyl)ethylene, or/?,//-dichlorodiphenyldichloroethylene
(p,p'-DDE), in its pure form is a white, crystalline solid, soluble in fats and most organic
solvents, but practically insoluble in water. DDE was an impurity (up to 4%) in commercial
DDT pesticide formulations and never has been produced commercially (NCI, 1978). DDE is
also produced as a break-down product of DDT by the environmental degradation or
metabolism.
The DDE that was used in the NCI (1978) studies was purchased from Aldrich Chemical
Company and was nearly purep,p-DDE isomer based on its melting point. Many of the recently
published studies of DDE also purchased the chemical from the Aldrich Chemical Company and
reported the purity as >99% (You et al., 1999a,b). For the remainder of this document, the
abbreviation DDE will refer to the/>,p'-DDE isomer. In cases where a second isomer was
utilized (0,/>-DDE), the differing orientation of the chlorines in the molecule will be specified.
The chemical structure ofp,//-dichlorodiphenyldichloroethylene is shown above (Figure 2-1).
Its physical and chemical properties, and other reference information are listed in Table 2-1.
Figure 2-1 Chemical Structure of /7,/7'-Dichlorodiphenyldichloroethylene
ci
Source: Chemfinder (2004)
DDE —January, 2008
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Table 2-1 Chemical and Physical Properties of ff,ff'-Dichlorodiphenyldichloroethylene
Property
Chemical Abstracts Registry
(CAS) No.
EPA Pesticide Chemical Code
Synonyms
Registered Trade Name(s)
Chemical Formula
Molecular Weight
Physical State
Boiling Point
Melting Point
Density (at 20 °C)
Vapor Pressure:
At 20 °C
At 25 °C
Partition Coefficients:
Log Kow
LogKoc
Solubility in:
Water
Other Solvents
Conversion Factors
(at 25 °C, 1 atm)
Information
72-55-9
NA
4,4'-DDE
p,p '-DDE
DDE
2,2-bis(4-chlorophenyl)-l,l-dichloroethene
2,2-bis(p-chlorophenyl)-l,l-dichloroethylene
DDT Dehydrochloride
l,l-dichloro-2,2-bis(p-chlorophenyl)ethylene
Dichlorodiphenyldichloroethylene
Dichlorodiphenyldichloroethylene, P,p'-
l,l'-dichloroethenylidene)bis(4-chlorobenzene)
Ethylene, 1 , 1 -dichloro-2,2-bis(p-chlorophenyl)-
No data
C14H8C14
318.03
Crystalline solid
336 °C
80-94 °C
No data
No data
6.0xlO-6mmHg
6.956 (Range of published literature: 4.88-7.2)
4.70
0.12mg/Lat25°C
(Range of publ. literature: 0.0011-0.12 mg/L)
lipids and most organic solvents
1 ppm= 13.01 mg/m3
1 mg/m3= 0.077 ppm
CmnYWcV ATCTYD nc\c\i\- tTCTYR <">nn/iv TT c THDA ci o««v»v TTcnc Connie
DDE —January, 2008
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3.0 USES AND ENVIRONMENTAL FATE
3.1 Production and Use
Currently, DDE is not commercially produced in the United States and has no
commercial use. It is a degradation product and impurity of the once commercially produced
pesticide, DDT (HSDB, 2004). DDE is the product of the dehydrohalogenation of DDT (loss of
one molecule of hydrochloric acid). DDT was once widely used in the U.S. as a broad spectrum
organochlorine pesticide to control insects in agriculture and insects that carry diseases such as
malaria and typhus (Gianessi and Puffer, 1992). Production of DDT in the United States was at
its peak in 1962 when 85,000 tons of chemical were produced and 334 agricultural products
using DDT were registered (Metcalf, 1995). DDT production in the United States declined from
82 million kg in 1962 to 2 million kg in 1971. On January 1, 1973, all uses of DDT in the
United States were canceled, except for emergency public health uses and a few other uses
permitted on a case-by-case basis. In smaller quantities, DDT production for export continued as
late as the 1980s (ATSDR, 2002; HSDB, 2004). Currently, DDT is not produced commercially
in the United States, but is still used in other countries. Until recently, DDT was produced and
used in Mexico; however, the production was ended in 1997 and use was phased-out by 2000
under the North American Agreement on Environmental Cooperation (Meister and Sine, 1999;
CEC, 2003). Analytical studies have revealed that DDE may be a contaminant in technical grade
insecticide dicofol (Risebrough et al., 1986). In addition, DDE can be a product of the
degradation of l,l,2,2-tetrachloro-2,2-bis(p-chlorophenyl)ethane, another DDT-related impurity
in dicofol (ATSDR, 2002).
3.2 Environmental Release
DDE is found in the environment as a result of contamination or breakdown of DDT.
Although DDT is no longer produced or used in the United States, the insecticide is still used in
other parts of the world. DDE currently found in the environment may be persistent residues
from earlier use of DDT or may be recently deposited following long-range atmospheric
transport from areas where DDT is still released to the environment. DDT that has entered the
atmosphere via spraying or volatilization can contaminate soils and surface waters by both wet
and dry deposition. In soil, DDT biodegrades to DDE under unflooded (generally aerobic)
conditions and to ODD (dichlorodiphenyldichloroethane) under flooded (generally anaerobic)
conditions (ATSDR, 2002).
Among the 1613 hazardous waste sites in the United States and its territories that have
been considered as candidates for inclusion in EPA's National Priorities List (NPL), at least 441
are known to be contaminated with DDT, DDE, and/or DDD. p,p'-DDE was found at 219 of
these sites. While not specifically targeted, o,p'-DDE was also present in at least four sites. Of
the 441 hazardous waste sites in which DDT, DDE, or DDD was detected, the contaminants
were identified in air samples at 32 sites, in surface water samples at 101 sites, in ground water
samples at 247 sites, and in sediment samples at 305 sites (HazDat, 2002).
DDE — January, 2008 3-1
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3.3 Environmental Fate
Air
During its period of agricultural use, DDT was sprayed primarily onto crops, forest lands,
surface water, and residential areas. Since DDE is an impurity and degradation product of DDT,
application of DDT led to the presence of DDE to the environment. In the atmosphere, DDE
may exist in the vapor phase, based on the vapor pressure of 6xlO"6 mm Hg at 25°C, or it can be
adsorbed onto particulate matter after release into the atmosphere (HSDB, 2004). In the
atmosphere, vapor-phase DDE reacts with photochemically-produced hydroxyl radicals and the
estimated half-life ranges from 17 hours to 2 days (Meylan and Howard, 1993). The longer half-
life was calculated from its estimated rate constant of 7.4xlO"12 cm3/molecule-sec at 25°C as
determined from a structure estimation method (Meylan and Howard, 1993). The reported
atmospheric half-life in sunlight at 40° latitude was calculated to range from 0.9 days in summer
to 6.1 days in winter (Callahan et al., 1979; Zepp and Cline, 1977). DDE is expected to undergo
direct photolysis since this compound absorbs light at wave lengths greater than 290 nm. Freitag
et al. (1979) observed 20% photomineralization of DDE when it was applied to a silica gel
irradiated with UV radiation (wavelengths >290 nm) for 7 days.
Particulate-phase DDE may be removed from the air by wet and dry deposition. Particle-
phase atmospheric DDE may be subject to long-range transport, which may be responsible for
the DDE present at sites distant from where DDT was applied. DDE's vapor pressure of 6x10"6
mm Hg (Bidleman, 1984) and Henry's Law constant of 4.16xlO"5 atm-m3/mole (Altschuh, 1999)
predict that it will volatilize from moist soils and surface waters; thus, DDE in the atmosphere
can be further redistributed by wet or dry deposition throughout the world. This process can be
repeated multiple times with volatilization in warm climates and deposition into cooler climates,
which is called global distillation. DDE volatilization is expected to be attenuated by adsorption
to carbon sources since DDE has high adsorption coefficients, i.e., log Koc values of 4.42 (Ding
and Wu, 1995), 4.70 (Sabljic, 1984), and 4.88 (Ding and Wu, 1997). Ding and Wu (1995, 1997)
measured empirical Koc values from soil column batches from Taichung, Taiwan. Sabljic (1984)
used structural modeling to predict the adsorption coefficient.
Terrestrial
Primarily due to high partition coefficients, DDE, is expected to adsorb strongly onto soil
particles. As a result of DDE's strong binding to soil, it will remain mostly on the surface layers
of soil (top 1.5-cm layer); small amounts leach into the lower soil layers and groundwater
(Callahan et al., 1979). In the process of weathering or aging, DDE becomes sequestered into
micropores of soil or sediment, decreasing its bioavailability (ATSDR, 2002). DDE has a low
water solubility (0.0011-0.12 mg/L) suggesting limited mobility in soils. In moist soils DDE
may volatilize based on the Henry's Law constant, but it is not expected to volatilize from dry
soils based on its vapor pressure (HSDB, 2004). Loss of DDE from dryer soil is primarily due to
the transport of the particulate mater to which the compound is bound.
Soil dissipation of DDE is much greater in tropical than in temperate regions. In
Alabama, a fugacity-based, multilayered soil exchange model predicted that 200-600 kg of DDE
is released into the air each year from a 1.23xlOn m2 area (the size of Alabama) that has a
DDE — January, 2008 3-2
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geometric mean of 7.9 ng/g dry weight of soil (23 ng/g dry weight of soil arithmetic mean),
which is -50% of the air burden (Harner et al., 2001). Reported half-life in the soil is 1000 days
(Mackay et al., 1997). Recent studies report the half-life of 2DDT (i.e., combined/?,/>'-DDT,
DDD, and DDE) and DDE in tropical regions to be between 22 (in Sudan) to 327 (in China) and
151 to 271 days (in Brazil), respectively. The DDE half-lives in temperate regions ranged from
837 to 6,087 days. The half-life of these compounds was also higher in acidic soils (half-life
>672 days, pH=4.5), and highest residue levels were found in muck soils and in deeply plowed,
unflooded fields (ATSDR, 2002).
Abiotic degradation
DDE is known to undergo photooxidation reactions on the surfaces of soil or sediment
(Baker and Applegate, 1970; Lichtenstein and Schultz, 1959; Miller and Zepp, 1979). The
conversion of DDT to DDE in soil was enhanced by exposure to sunlight. In a 90-day
experiment, 91% of the initial concentration of DDT remained in the soil when not exposed to
light (i.e., a dark control); 65% remained for the sample exposed to light (Racke et al., 1997).
No information was located on the abiotic mineralization of DDE; UV-irradiation of uC-p,p' -
DDT on soil for 10 hours mineralized less than 0.1% of the initial amount (Vollner and Klotz
1994).
Biotic degradation
DDT biodegrades primarily to DDE under aerobic anaerobic conditions due to soil
microorganisms including bacteria, fungi, and algae (Arisoy, 1998; U.S. EPA, 1979;
Lichtenstein and Schulz, 1959; Menzie, 1980; Stewart and Chisholm, 1971; Verma and Pillai,
1991). Mineralization of DDT and DDE was observed in laboratory experiments using
Phanerochaete chrysosporium (a white rot fungus) (Aislabie et al., 1997; Singh et al., 1999).
Other soil microorganisms, such asAerobacter aerogenes, Pseudomonasfluorescens, E. coli,
and Klebsiellapneumoniae, have also been shown to have the capability to degrade DDT under
both aerobic and anaerobic conditions, forming 4-chlorobenzoic acid and DDE, respectively
(Singh et al., 1999). Beyond these limited examples, DDE is often resistant to biodegradation
under aerobic and anaerobic conditions (Strompl and Thiele, 1997).
DDT breaks down into DDE and DDD in soil, and the parent-to-metabolite ratio (DDT to
DDE or DDD) decreases with time due to weathering of DDT and its metabolites into soil
micropores. The parent-to-metabolite ratio may vary considerably with soil type. DDT was
much more persistent in muck soils than in dry forest soils. A study of agricultural soils in the
corn belt of the central United States found the ratio of p,p' -DDT//?,/?' -DDE varied from 0.5 to
6.6 with three-quarters of the soils having ratios above 1 (Aigner et al., 1998). In a study of
forest soils in Maine, the half-life for the disappearance of DDT residues was noted to be 20-30
years (Dimond and Owen, 1996). A study of DDT in agricultural soils in British Colombia,
Canada reported that over a 19-year period, there was a 70% reduction of DDT in muck soils and
a virtual disappearance of DDT from loamy sand soils (Aigner et al., 1998).
Aquatic
DDE may revolatilize if deposited into water; however, it preferentially adsorbs to
particulate matter and partitions into the sediment as predicted by the high partitioning
DDE — January, 2008 3-3
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coefficients. The average log Koc value of DDE in lake sediment was 4.58 (van den Hoop, et al.,
1999). Lyman et al. (1990) estimated the volatilization half-life of DDE from a model river and
model lake to be 2 and 18 days, respectively. When adsorption was considered in a
volatilization model performed by EPA's EXAMS II Computer Simulator (U.S. EPA, 1987) the
half-life was determined to be 5 years. Hydrolysis is not a major source of degradation;
however, photolysis is possible on sunlit water surfaces. The half-life of DDE in irradiated
water (310-410 nm) due to photolysis was 15 and 26 hours (Draper, 1985). DDE did not display
any sign of degradation after 12 months when exposed to seawater and sediment (Vind et al.,
1973).
In laboratory experiments with marine sediment, DDE was shown to be dechlorinated to
DDMU (l-chloro-2,2-bis[p-chlorophenyl]ethylene) under methanogenic or sulfidogenic
conditions (Quensen et al., 1998). The rate of DDE dechlorination to DDMU was found to be
dependent on the presence of sulfate and temperature (Quensen et al., 2001). DDMU degrades
further under anaerobic conditions to 2,2-bis(chlorophenyl)acetonitrile (DDNU) and other
subsequent degradation species, such as 2,2-bis(chlorophenyl)ethanol (DDOH) and
2,2-bis(chlorophenyl)acetic acid (DDA), through chemical action (Heberer and Diinnbier, 1999;
Wareetal., 1980).
DDE can bioaccumulate due to its high lipophilicity and long half-life. It biomagnifies
up the food chain, i.e., concentrations progressively increase in the tissues of plants and animals
in successive trophic levels. It also bioconcentrates in aquatic organisms and bioaccumulates in
the food chain (ATSDR, 2002). A field study was conducted where DDE was introduced into a
flooded quarry and the water, sediment, and biota were monitored for DDE levels for one year
(Callahan et al., 1979). The equilibrium of DDE between water and zooplankton was attained
within a day, with a bioconcentration factor (BCF) of 3-6xlO+4. DDE reached equilibrium
between water and bluegill, a resident quarry fish, after 60 days, with a BCF of about l.lxlO+5;
the BCF for trout in this study was obtained after 108 days, and was 1.8xlO+5 (Callahan et al.,
1979). Callahan et al. (1979) reported the findings from a terrestrial-aquatic microcosm
experiment on the fate of 3.8 ppb DDE in water. The BCFs were calculated to be 3.6xlO+4,
5.9xlO+4, 1.2xlO+4, and l.lxlO+4 for snail, mosquito larvae, fish, and algae respectively.
DDE BCF values have also been reported in the following aquatic organisms (HSDB,
2004):
• Rainbow trout: 81,000
• Flathead minnow: 51,000
• Fish (Species not reported; static microcosm study): 27,500
• Gambusia affmis (3 days): 217
• Zooplankton: 28,600
DDE — January, 2008 3-4
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3.4 Summary
DDE is an impurity and a degradation product of DDT, a pesticide that was manufactured
and used in the US, and is still manufactured and used in other parts of the world. Thus, DDE
has been released to the environment as a result of the use of DDT as an insecticide. In the air,
DDE can exist in both the vapor and paniculate phases. Vapor-phase DDE may be degraded by
reaction with photochemically-produced hydroxyl radicals; it also may undergo direct
photolysis. Particulate-phase DDE may undergo long-distance atmospheric transport and can be
removed from the atmosphere by wet and dry deposition.
In the soil, DDE is expected to have limited mobility and bioavailability because it binds
to soil particles. Volatilization from moist soil surfaces is expected to be an important
environmental fate process; however, adsorption to the soil may attenuate this process. DDE is
not expected to volatilize from dry soil surfaces based on its vapor pressure. DDE is resistant to
degradation in the soil, and the primary loss of DDE may be due to erosion of the particles to
which the compound is bound.
In water, volatilization from water surfaces is expected to be an important environmental
fate process, based upon its Henry's Law constant. However, volatilization from water surfaces
is expected to be attenuated severely by adsorption to suspended solids and sediment in the water
column. Biodegradation of DDE in water is expected to be very slow. DDE bioconcentrates and
bioacumulates in the aquatic environment; BCF values of up to 180,000 have been reported in
fish, suggesting that bioconcentration in aquatic organisms is very high.
DDE — January, 2008 3-5
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DDE — January, 2008 3-6
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4.0 EXPOSURE FROM DRINKING WATER
4.1 Introduction
EPA used data from several sources to evaluate the potential for occurrence of DDE in
Public Water Systems (PWSs). The primary source of drinking water occurrence data for DDE
was the first Unregulated Contaminant Monitoring Regulation (UCMR 1) program. The Agency
also evaluated ambient water quality data from the United States Geological Survey (USGS).
4.2 Ambient Occurrence
4.2.1 Data Sources and Methods
USGS instituted the National Water Quality Assessment (NAWQA) program in 1991 to
examine ambient water quality status and trends in the United States. NAWQA is designed to
apply nationally consistent methods to provide a consistent basis for comparisons among study
basins across the country and over time. These occurrence assessments serve to facilitate
interpretation of natural and anthropogenic factors affecting national water quality. For more
detailed information on the NAWQA program design and implementation, please refer to Leahy
and Thompson (1994) and Hamilton and colleagues (2004).
Study Unit Monitoring
The NAWQA program conducts monitoring and water quality assessments in significant
watersheds and aquifers referred to as "study units." NAWQA's sampling approach is not
"statistically" designed (i.e., it does not involve random sampling), but it provides a
representative view of the nation's waters in its coverage and scope. Together, the 51 study units
monitored between 1991 and 2001 include the aquifers and watersheds that supply more than
60% of the nation's drinking water and water used for agriculture and industry (NRC, 2002).
NAWQA monitors the occurrence of chemicals such as pesticides, nutrients, volatile organic
compounds (VOCs), trace elements, and radionuclides, and the condition of aquatic habitats and
fish, insects, and algal communities (Hamilton et al., 2004).
Monitoring of study units occurs in stages. Between 1991 and 2001, approximately one-
third of the study units at a time were studied intensively for a period of three to five years,
alternating with a period of less intensive research and monitoring that lasted between five and
seven years. Thus all participating study units rotated through intensive assessment in a ten-year
cycle (Leahy and Thompson, 1994). The first ten-year cycle was called "Cycle 1." Summary
reports are available for the 51 study units that underwent intensive monitoring in Cycle 1
(USGS, 2001a). Cycle 2 monitoring is scheduled to proceed in 42 study units from 2002 to 2012
(Hamilton et al., 2004).
DDE — January, 2008 4-1
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Pesticide National Synthesis
Through a series of National Synthesis efforts, the USGS NAWQA program is preparing
comprehensive analyses of data on topics of particular concern. These data are aggregated from
the individual study units and other sources to provide a national overview.
The Pesticide National Synthesis began in 1991. Results from the most recent USGS
Pesticide National Synthesis analysis, based on complete Cycle 1 (1991-2001) data from
NAWQA study units, are posted on the NAWQA Pesticide National Synthesis website (Martin
et al., 2003; Kolpin and Martin, 2003; Nowell, 2003; Nowell and Capel, 2003). USGS considers
these results to be provisional. Data for surface water, ground water, bed sediment, and biota are
presented separately, and results in each category are subdivided by land use category. Land use
categories include agricultural, urban, mixed (deeper aquifers of regional extent in the case of
ground water), and undeveloped. The National Synthesis analysis for pesticides is a first step
toward the USGS goal of describing the occurrence of pesticides in relation to different land use
and land management patterns, and developing a deeper understanding of the relationship
between spatial occurrence of contaminants and their fate, transport, persistence, and mobility
characteristics.
The surface water summary data presented by USGS in the Pesticide National Synthesis
(Martin et al., 2003) only include stream data. Sampling data from a single one-year period,
generally the year with the most complete data, were used to represent each stream site. Sites
with few data or significant gaps were excluded from the analysis. NAWQA stream sites were
sampled repeatedly throughout the year to capture and characterize seasonal and hydrologic
variability. In the National Synthesis analysis, the data were time-weighted to provide an
estimate of the annual frequency of detection and occurrence at a given concentration.
The USGS Pesticide National Synthesis only analyzed ground water data from wells;
data from springs and agricultural tile drains were not included. The sampling regimen used for
wells was different than that for surface water. In the National Synthesis analysis (Kolpin and
Martin, 2003), USGS uses a single sample to represent each well, generally the earliest sample
with complete data for the full suite of analytes.
NAWQA monitored bed sediment and fish tissue at sites considered likely to be
contaminated and sites that represent various land uses within each study unit. Most sites were
sampled once in each medium. In the case of sites sampled more than once, a single sample was
chosen to represent the site in the Pesticide National Synthesis analysis (Nowell, 2003). In the
case of multiple bed sediment samples, the earliest one with complete data for key analytes was
used to represent the site. In the case of multiple tissue samples, the earliest sample from the
first year of sampling that came from the most commonly sampled type offish in the study unit
was selected.
As part of the National Pesticide Synthesis, USGS also analyzed the occurrence of select
semivolatile organic compounds (SVOCs) in bed sediment at sites considered likely to be
contaminated and sites that represent various land uses within each study unit (Nowell and
DDE — January, 2008 4-2
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Capel, 2003). Most sites were sampled only once. When multiple samples were taken, the
earliest one was used to represent the site in the analysis.
Over the course of Cycle 1 (1991-2001), NAWQA analytical methods may have been
improved or changed. Hence, reporting levels (RLs) varied over time for some compounds. In
the summary tables, the highest RL for each analyte is presented for general perspective. In the
ground water, bed sediment, and tissue data analyses, the method of calculating concentration
percentiles sometimes varied depending on how much of the data was censored at particular
levels by the laboratory (i.e., because of the relatively large number of non-detections in these
media).
4.2.2 Results
Surface Water and Ground Water
Under the NAWQA program, USGS monitored DDE (specifically/\p -DDE, the most
common isomer) between 1992 and 2001 in representative watersheds and aquifers across the
country. Reporting limits in surface water and ground water varied but did not exceed 0.006
|ig/L. Results for surface water and ground water are presented in Tables 4-1 and 4-2.
Table 4-1 USGS Pesticide National Synthesis Summary of NAWQA Monitoring ofp,p'-
DDE in Ambient Surface Water, 1992-2001
Land Use Type
Agricultural
Mixed
Undeveloped
Urban
No. of Samples
(and No. of
Sites)
1,885 (78)
1,021 (47)
60(4)
900 (33)
Detection
Frequency
4.84%
6.14%
3.66%
1.68%
50th Percentile
(Median)
Concentration
-------
Table 4-2 USGS Pesticide National Synthesis Summary of NAWQA Monitoring of p,p '-
DDE in Ambient Ground Water, 1992-2001
Land Use Type
Agricultural
Mixed (Major
Aquifer)
Undeveloped
Urban
No. of Wells
1,443
2,716
67
834
Detection
Frequency
3.26%
2.65%
7.46%
3.96%
50th Percentile
(Median)
Concentration
-------
Table 4-3 USGS Pesticide National Synthesis Summary of NAWQA Monitoring of p,p '-
DDE in Bed Sediment, 1992-2001
Land Use Type
Agricultural
Mixed
Undeveloped
Urban
No. of Sites
282
338
224
166
Detection
Frequency in
samples
48%
46%
22%
70%
50th Percentile
(Median)
Concentration
0.98 ug/kg
0.81 ug/kg
-------
NAWQA data indicate thatp,p'-DDE occurs in fish tissue (Table 4-4) at detection
frequencies ranging from 44% of samples in undeveloped settings to 89% in agricultural
settings, 89% in urban settings, and 93% in mixed land use settings. The 95th percentile
concentrations in fish tissue were found to range from 128 |ig/kg wet weight (undeveloped
settings) to 2180 |ig/kg wet weight (agricultural settings). The highest concentration, 7300
Hg/kg wet weight, was found in an agricultural setting (Nowell, 2003).
Results of monitoring for o,p'-DDE in bed sediment and fish tissue are presented in
Tables 4-5 and 4-6.
Table 4-5 USGS Pesticide National Synthesis Summary of NAWQA Monitoring of o,p '-
DDE in Bed Sediment, 1992-2001
Land Use Type
Agricultural
Mixed
Undeveloped
Urban
No. of Sites
278
327
221
164
Detection
Frequency in
samples
2.6%
1.6%
0.0%
3.7%
50th Percentile
(Median)
Concentration
-------
Table 4-6 USGS Pesticide National Synthesis Summary of NAWQA Monitoring of o,p '-
DDE in Whole Fish, 1992-2001
Land Use Type
Agricultural
Mixed
Undeveloped
Urban
No. of Sites
204
206
162
99
Detection
Frequency
7.0%
3.2%
0.0%
6.4%
50th Percentile
(Median)
Concentration
-------
year, six-month interval monitoring schedule for ground water systems. Although UCMR 1
monitoring was conducted primarily between 2001 and 2003, some results were not collected
and reported until as late as 2006.
The objective of the UCMR 1 sampling approach for small systems was to collect
contaminant occurrence data from a statistically selected, nationally representative sample of
small systems. The small system sample was stratified and population-weighted, and included
some other sampling adjustments such as allocating a selection of at least two systems from each
state. With contaminant monitoring data from all large PWSs and a statistical, nationally
representative sample of small PWSs, the UCMR 1 List 1 Assessment Monitoring program
provides a contaminant occurrence data set suitable for national drinking water estimates.
4.3.2 Derivation of the Health Reference Level
To evaluate the systems and populations exposed to DDE through PWSs, the monitoring
data were analyzed against the Minimum Reporting Level (MRL) and a benchmark value for
health that is termed the Health Reference Level (HRL). Two different approaches were used to
derive the HRL, one for chemicals that cause cancer and exhibit a linear response to dose and the
other applies to noncarcinogens and carcinogens evaluated using a non-linear approach.
The HRL of 0.2 |ig/L for DDE considers the potential carcinogenic effects of DDE (see
Section 8.2.4). The HRL is based on the occurrence of liver tumors in mice following chronic
exposures (NCI, 1978; please see the Dose-Response Section of this document for more
information). No RfD is currently available for DDE.
4.3.3 Results
As a List 1 contaminant, 4,4'-DDE (as thep,p'-DDE isomer is also known) was scheduled
to be monitored by all large CWSs and NTNCWSs and a statistically representative sample of
small CWSs and NTNCWSs. The data presented in this report reflect UCMR 1 analytical
samples submitted and quality-checked under the regulation as of March 2006. 4,4'-DDE data
were collected and submitted by 797 (99.6 percent) of the 800 small systems selected for the
small system sample and 3077 (99.3 percent) of the 3100 large systems defined as eligible for
the UCMR 1 large system census. 4,4'-DDE data have been analyzed at the level of simple
detections (at or above the minimum reporting level, >MRL, or >0.8 |ig/L). Since the health
reference level (0.2 |ig/L) is less than the MRL, the data are not analyzed at the level of the HRL
or half the HRL.
EPA set the MRL for UCMR 1 contaminants based on the capability of analytical
methods, not anticipated health levels. For many UCMR 1 contaminants, including DDE, the
MRL was determined by multiplying by 10 the least sensitive method's minimum detection
limit, or, when available, multiplying by 5 the least sensitive method's estimated detection limit
(U.S. EPA, 2000e). MRLs were set approximately an order of magnitude higher than detection
limits to ensure consistency, accuracy, and reproducibility of results.
DDE — January, 2008 4-8
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For the 1999 UCMR 1, EPA approved three analytical methods for the analysis of DDE.
These included EPA Methods 508, 508.1 and 525.2. In setting the minimum reporting limit
(MRL) for UCMR 1 contaminants, EPA chose the method detection limit (MDL) for the least
sensitive method and multiplied this MDL by a factor of 10. Of the three methods approved for
DDE analysis, method 525.2 is the least sensitive. Data from four MDL studies are included in
Method 525.2, extraction using either C-18 disk or cartridge with the subsequent analyses
performed using either a quadrupole or ion trap mass spectormeter. The MDLs generated using
these four options were 0.054, 0.070, 0.070 and 0.075 |ig/L. This resulted in an MRL of 0.80
|ig/L (i.e., 10 times the MDL of 0.075 |ig/L = 0.75 |ig/L and then rounded to 0.80 |ig/L).
Results of the analysis are presented in Tables 4-7 and 4-8. No detections of 4,4'-DDE
were found in any samples from small systems. DDE was detected at a single large system; this
ground water system represented 0.03% of large PWSs and 0.01% of the population served by
them (approximately 18,000 people). The concentration of the single detection was 3 |ig/L.
DDE — January, 2008 4-9
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Table 4-7 Summary UCMR 1 Occurrence Statistics for 4,4'-DDE in Small Systems
(Based on Statistically Representative National Sample of Small Systems)
Frequency Factors
Total Number of Samples
Percent of Samples with Detections
99 Percentile Concentration (all samples)
Health Reference Level (HRL)
Minimum Reporting Level (MRL)
Maximum Concentration of Detections
99th Percentile Concentration of Detections
Median Concentration of Detections
Total Number of PWSs
Number of GW PWSs
Number of SW PWSs
Total Population
Population of GW PWSs
Population of SW PWSs
Occurrence by System
PWSs (GW & SW) with Detections (> MRL)
Occurrence by Population Served
Population Served by PWSs with Detections
UCMR Data -
Small Systems
3,251
0.00%
ViHRL, or PWSs > HRL = PWSs with at least one sampling result greater than or
equal to the MRL, exceeding the ViHRL benchmark, or exceeding the HRL benchmark, respectively; Population Served by
PWSs with detections, by PWSs >1/2HRL, or by PWSs >HRL = population served by PWSs with at least one sampling result
greater than or equal to the MRL, exceeding the 'AHRL benchmark, or exceeding the HRL benchmark, respectively.
Notes:
-Small systems are those that serve 10,000 persons or fewer.
-Only results at or above the MRL were reported as detections. Concentrations below the MRL are considered non-detects.
-Due to differences between the ratio of GW and SW systems with monitoring results and the national ratio, extrapolated GW
and SW figures might not add up to extrapolated totals.
-The HRL used in this analysis is a draft value for working review only.
DDE —January, 2008
4-10
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Table 4-8 Summary UCMR 1 Occurrence Statistics for 4,4'-DDE in Large Systems
(Based on the Census of Large Systems)
Frequency Factors
Total Number of Samples
Percent of Samples with Detections
99 Percentile Concentration (all samples)
Health Reference Level (HRL)
Minimum Reporting Level (MRL)
Maximum Concentration of Detections
99 Percentile Concentration of Detections
Median Concentration of Detections
Total Number of PWSs
Number of GW PWSs
Number of SW PWSs
Total Population
Population of GW PWSs
Population of SW PWSs
Occurrence by System
PWSs (GW & SW) with Detections (> MRL)
GW PWSs with Detections
SW PWSs with Detections
Occurrence by Population Served
Population Served by PWSs with Detections
Pop. Served by GW PWSs with Detections
Pop. Served by SW PWSs with Detections
UCMR Data -
Large Systems
30,546
0.003%
ViHRL, or PWSs > HRL = PWSs with at least one sampling result greater than or
equal to the MRL, exceeding the VzHRL benchmark, or exceeding the HRL benchmark, respectively; Population Served by
PWSs with detections, by PWSs >1/2HRL, or by PWSs >HRL = population served by PWSs with at least one sampling result
greater than or equal to the MRL, exceeding the 'AHRL benchmark, or exceeding the HRL benchmark, respectively.
Notes:
-Large systems are those that serve more than 10,000 persons.
-Only results at or above the MRL were reported as detections. Concentrations below the MRL are considered non-detects.
-The HRL used in this analysis is a draft value for working review only.
DDE —January, 2008
4-11
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Regional Patterns
DDE was only detected in one sample in all of the UCMR 1 sampling. This single
detection was in a ground water sample taken in the state of Alabama (see Figure 4-1). Since
only one system detected the contaminant, no further spatial analysis of this contaminant is
presented.
Figure 4-1 Geographic Distribution of 4,4'-DDE in UCMR 1 Monitoring - States With
at Least One Detection at or Above the MRL (0.8 ug/L)
DCD
D Guam D Mariana Is.
D Virgin Is. D Puerto Rico
D Tribes
Entities with No Detections
Entities with Detections (> 0.8 ug/L)
4.4 Summary
The USGS NAWQA program conducted monitoring of/?,/?'-DDE in ambient surface
water, ground water, bed sediment, and fish tissue from 1992 to 2001 (Cycle 1). Four types of
land use were considered in this monitoring effort: agricultural, mixed, undeveloped, and urban.
Reporting limits varied over the course of the cycle, but did not exceed 0.006 jig/L in water, 1
Hg/kg (dry weight) in bed sediment, and 5 |ig/kg (wet weight) in whole fish. Summary results
are reported in the USGS National Pesticide Synthesis. In none of the 3866 ambient surface
water samples or 5060 ambient ground water samples was DDE found at concentrations that
exceeded the URL of 0.2 |ig/L. The highest concentrations of/?,/?'-DDE in surface water (0.062
|ig/L), ground water (0.008 |ig/L) were observed in areas of agricultural land use; however, the
highest concentration of/?,/?'-DDE in bed sediment (440 |ig/kg dry weight) was observed in an
area of mixed land use. Median concentrations in bed sediment ranged from less than the
reporting limit to 2.15 |ig/kg dry weight in the different land use settings. The median
concentrations in surface water and ground water were below the reporting limit. In bed
DDE —January, 2008
4-12
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sediment, it was found most frequently in urban areas (70% of samples). In ground water it was
found most frequently in undeveloped areas (7.5% of samples), and in surface water it was found
most frequently in areas of mixed land use (6.1% of samples). The NAWQA program also
monitored the less common isomer o,p'-DDE in bed sediment and fish tissue. The highest
concentration of o,p'-DDE in bed sediments (26.7 |ig/kg dry weight) was found in an urban
setting. In bed sediments, o,p'-DDE was detected most frequently in urban areas (3.7% of
samples).
Under the UCMR 1, DDE was monitored in 33,797 finished drinking water samples from
3077 large and 797 small public water systems in the United States. At the MRL of 0.8 |ig/L, no
detections of DDE were found in any of the samples from small systems. DDE was detected at a
single large system; this ground water system represented 0.03% of large PWSs and 0.01% of
the population served by them (approximately 18,000 people). Since the HRL (0.2 |ig/L) is less
than the MRL (0.8 |ig/L), the data are not analyzed at the level of the HRL or half the HRL.
DDE — January, 2008 4-13
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DDE — January, 2008 4-14
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5.0 EXPOSURE FROM MEDIA OTHER THAN WATER
5.1 Exposure from Food
Currently, food provides the primary source of DDE exposure to the general population
(ATSDR, 2002). Although DDT is no longer used in the US, it still is used for pest control in
some areas of the world (ATSDR, 2002), and depending on use and importation of foods from
other countries, there may be some dietary exposure to DDE (Coulston, 1985). DDE is more
persistent than DDT and may be expected to remain in animal and human tissues and in
commodities for a notably longer time (Arifio, 1995). In addition, DDE bioaccumulates in the
environment through terrestrial and aquatic food chains, creating a potential for dietary
exposures far greater than environmental levels. Research has found that higher amounts of
DDT are consumed by people eating fish from the Great Lakes (Hanrahan et al., 1999; Laden et
al., 1999), and that the populations at greatest risk to DDT are the indigenous people in the
Arctic who eat traditional foods (e.g., seals, caribou, narwhals, etc.) (Kuhnlein et al., 1995).
As in other media, DDT levels in food generally are declining. Data taken from the
Market Basket Surveys between 1965 and 1975 showed an 86% decline in DDT levels in all
kinds of food (U.S. EPA, 1980); these studies found decreased DDE levels in all classes of food
tested. Detectable quantities of DDE are still present in many commodities. The U.S. FDA's
analysis of 13,283 imported agricultural commodities for pesticide residues (1981 to 1986)
showed that although no commodities exceeded the EPA tolerance levels for DDT or DDE,
3.1% of the samples had detectable levels of DDT or DDE (Hundley et al., 1988). DDE residues
also were found in 41 out of the 6,970 produce samples (0.6%) tested in a 1989 pesticide
screening program of produce delivered to supermarkets in Texas (Schattenberg and Hsu, 1992).
5.1.1 Concentration in Non-Fish Food Items
The most recent FDA market basket study summary (1991-2001) found DDE in 157 of
279 food types analyzed (U.S. FDA, 2003). They reported that mean DDE concentrations
ranged from 0.0001 ppm (several food types) to 0.0221 ppm (salted butter). The highest mean
DDE concentrations were identified in salted butter (0.0221 ppm, n=36), boiled spinach (0.0109
ppm, n=36), baked salmon (0.0079 ppm, n = 16), American cheese (0.0067 ppm, n=36), pan-
cooked lamb chop (0.0055 ppm, n=34), and boiled collards (0.0055, n=34). The highest
individual sample concentrations of DDE were in salted butter (0.1020 ppm), American cheese
(0.0480 ppm), boiled spinach (0.0370 ppm), pan-cooked pork sausage (0.0300 ppm), and pan-
cooked lamb chop (0.0300 ppm). Based on these data, it appears that the highest concentrations
of DDE can be found in dairy, meat, and leafy greens; fish will be discussed further in the
following section.
Data from other recent studies reporting only pesticide occurrences and not the
concentration levels are summarized in Table 5-1.
DDE — January, 2008 5-1
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Table 5-1 Percent of food samples with detectable DDE levels
Year(s)
1989-1991
1989
1988
1984-1986
1981-1986*
DDE
Isomer
p,p' -DDE
DDE
p,p' -DDE
DDE
p,p'-DDE
DDE
DDE
% of samples
with DDE
0.59%
0.99%
0.25%
1.5%
0.18%
23.1%
3.1%
Total No. of
Samples
6,970
13,085
13,085
13,980
13,980
1,872
13,283
Source
Schattenberg and Hsu, 1992
Minyard and Roberts, 1991
Minyard and Roberts, 1991
Minyard and Roberts, 1991
Minyard and Roberts 1991
Gunderson, 1995b
Hundley etal., 1988
*data on imported agricultural commodities
o,p'-DDE was not analyzed in the most recent U.S. Market Basket study; however, in a
previous report, this congener was detected only 8 times in 4 different food items at an average
concentration of 0.0025 |ig/g (whilep,p'-DDE was detected 1700 times in 142 different food
items at an average concentration of 0.0026 |ig/g; Kan- Do Office and Pesticide Team, 1995).
Another study of ready to eat foods (Schecter and Lingjun, 1997) measured DDE levels
in four types of popular U.S. fast foods. The chemical was detected in all four types of food with
concentrations of 3170 pg/g (ppt) in McDonald's Big Mac Hamburger, 650 pg/g in Pizza Hut's
Personal Pan Pizza Supreme, 180 pg/g in Kentucky Fried Chicken (KFC) three piece original
recipe mixed dark and white meat luncheon package, and 2780 pg/g in Haagen-Daz chocolate-
chip ice cream. These and previous data confirm that DDE bioconcentrates in meat and milk as
can be predicted from its high Koc. Thus, it is expected that those foods are potentially the
biggest source of the DDT metabolite in the human diet. Conventional and "organic" baby foods
from grocery stores in Michigan were evaluated for DDE and various other organochlorine
pesticide residues (Moore et al., 2000); however, none of the foods contained residues of
pesticides. The products tested were applesauce, pears, winter squash, and carrots.
DDE concentration was analyzed 129 samples of various Spanish meat products (Arifio
et al., 1995); the frequency of detection for DDE was between 78 and 100%. The mean levels
were below 10 ng/kg in all meat products, except for pork bologna where the mean
concentration measured was 16 ng/kg (total range: 4 to 30.1 ng/kg). In addition, this study
investigated the effect of commercial processing on the DDE residues and found DDE to be
resistant to degradation under the conditions of ham ripening, sausage cooking, and pork
bologna cooking. Another Spanish study supported the findings that the curing processes had
little or no effect on DDE (Bayarri et al.,1998); however, a common bacteria found in meat,
Micrococcus varians, reduced DDE concentrations by 17.7% in nutrient media. Measured levels
of DDE in sausages were as follows: salchion, 2.5-1.8 ng/g lipid; chorizo vela, 5.7-7.4 ng/g lipid;
and chorizo de Pamplono, 1.5-2.4 ng/g.
DDE —January, 2008
5-2
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Recent research conducted by the U.S. Department of Agriculture, through its Pesticide
Detection Program, found 17% of milk samples contained an average of 0.002 ppm DDE
(Benbrook, 2002). Another study noted that mean DDE levels in cow's milk in Southern
Ontario, Canada, have declined from 96 ng/g lipid in 1970-1971 to 16 ng/g lipid in 1985-1986
(Frank and Braun, 1989). Waliszewski et al. (1996) detected p,p'-DDE in 46% of the 192
samples of cow's milk collected between April and November of 1993 from random farms in the
central coastal region of Veracruz, Mexico; the highest concentration measured was 0.107
mg/kg, while the mean was 0.028 mg/kg.
Due to its high lipophilicity DDE selectively partitions into human fatty tissue and breast
milk. The fat content of milk fluctuates; therefore, it is more accurate to measure and express
DDE content on a lipid basis (i.e., ng/g lipid rather than |ig/mL milk), which has become the
standard for hydrophobic pollutants. These compounds are usually found in human breast milk
in concentrations higher than in cow's milk (human breast milk has a higher fat content than
cow's milk) or other infant foods. This leads to generally higher dietary exposure to DDE for
breast-fed infants. However, DDE concentrations in human breast milk have been declining
steadily across the U.S., Canada, and Western Europe (ATSDR, 2002).
Kalantzi et al. (2001) measuredp,p'-DDE and o,//-DDE concentrations in butter
throughout the world. The eighteen samples taken in the United States yielded an average of
24,070 pg/g lipid and a range of 1620-140,380 pg/g lipid forp,p'-DDE; o,p'-DDE was not
detected in any of those samples. Another study, focusing on DDE concentration in Spanish
cheeses, (Bentabol, 1995) reported the following mean (and range) values, estimated from 146
samples of various cheeses: 40.7 |lg/kg (4-354 |lg/kg) fat basis for/?,//-DDE; 6.9 |lg/kg (1-101
|ig/kg) fat basis for o,p'-DDE. DDE concentrations also were measured in Greek cheese
(Mallatou et al., 1996). The mean/?,//-DDE concentration in 28 samples was 37 ng/g fat, with a
range of 20-70 ng/g fat. The same study reported a mean p,p*'-DDE concentration from 38
samples of bovine milk from Greece to be 22 ng/g fat, with a range of 14-32 ng/g fat.
Mean DDE levels were reported for milk and cheese in Egypt. Abou-Arab (1997) reports
mean p,p*-DDE concentrations of 64 |ig/kg fat in milk and 4 |ig/kg fat in cheese as measured
from 25 samples collected from different regions. The respective mean o,p'-DDE concentrations
were 84 and 5 |ig/kg fat in milk and cheese.
5.1.2 Concentrations in Fish, Shellfish and Marine Mammals
The high lipid solubility of DDE combined with its extremely long half-life is
responsible for its high bioconcentration in aquatic organisms. This results in a progressive
biomagnification of DDE in organisms at the top of the food chain. It is possible that humans
may be the ultimate consumer of some of these organisms (ATSDR, 2002).
The most recent FDA market basket study summary (1991-2001), analyzed eight fish-
and shellfish-based food items for DDE, with mean concentrations ranging from 0.0008 to
0.0079 ppm (U.S. FDA, 2003).
DDE — January, 2008 5-3
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The USGS NAQWA program monitors water quality in more than 50 major river basins
and aquifer systems. A summary of their p,p'-DDE and o,p'-DDE data for whole fish tissue can
be found in the preceding section (Tables 4-4 and 4-6). These data constitute a national dataset
for DDE fish tissue concentrations; however, the types offish analyzed in this assessment may
not be the same types eaten by humans. Median concentrations in fish tissues ranged from 3.5 to
43.2 |ig/kg wet weight.
Giesy et al. (1994) measured the concentrations of the two DDE isomers in fish (8
species, 23 samples) from 3 rivers in Michigan between August and September 1990. They
reported ap,p'-DDE range of 3.54-627.13 |ig/kg wet weight, and o,p'-DDE range of 0.15-37.95
Hg/kg wet weight. The high ratios of DDE to DDT (5:1 to 758:1) suggested that the DDE
accumulation in the fish was a result of direct DDE exposure in the food chain rather than a
recent exposure to DDT.
There has been a marked decline in the levels of DDT and related compounds in fish,
shellfish, and aquatic mammals since the early 1970s (Addison and Stobo, 2001; Bard, 1999;
Lauenstein, 1995; Lieberg-Clark et al., 1995; Odsjo et al., 1997; Schmitt et al., 1990). The
National Contaminant Biomonitoring Program (Schmitt et al., 1990) reported that the mean
concentration ofp,p'-DDE has decreased from 260 |ig/kg in 1976 to 190 |ig/kg in 1984 (fish
sampled at 112 locations across the United States).
Recent studies reporting concentrations of DDT and its metabolites in various fish,
shellfish, and marine mammals are shown in Table 5-2. Only studies reporting DDE
concentrations have been included. Studies reporting combined concentration levels for DDT
and its metabolites have been omitted.
Table 5-2 Concentration of DDE in fish, shellfish, and marine mammals*
Species
Location
Year
Mean concentration, (range, if
applicable)
Reference
Fish and shellfish
Clams
Clams
Clams
Clams
Perch (n=5)
Perch (n=5)
Pike (n=5)
San Joaquin River
(Orestimba creek)
San Joaquin River (Dry
creek)
San Joaquin River
(Mokelumne River)
San Joaquin River
(Stanislaus River)
Lake 0rsj0en, Norway,
Mid-lake
Lake 0rsj0en, Norway
Lake 0rsj0en, Norway,
Mid-lake
1992
1992
1992
1992
1994
1994
1994
3,300 ng/g (w.w.)
25 ng/g (w.w.)
13 ng/g (w.w.)
22 ng/g (w.w.)
0.53 ng/g (w.w.), 757 ng/g (f.w.)
2.56 ng/g (w.w.), 5,120 ng/g
(f.w.)
3.5 ng/g (w.w.), 3,888 ng/g (f.w.)
Pereira et al.
(1996a)
Pereira et al.
(1996a)
Pereira et al.
(1996a)
Pereira et al.
(1996a)
Brevik et al.
(1996)
Brevik et al.
(1996)
Brevik et al.
(1996)
DDE —January, 2008
5-4
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Species
Lake trout (n=59)
Rainbow smelt
(n=8)
Location
Lake Ontario
Lake Ontario
Year
1992
1992
Mean concentration, (range, if
applicable)
1.159ug/g(w.w.)
0.256 ug/g (w.w.)
Reference
Kiriluk et al.
(1995)
Kiriluk et al.
(1995)
Marine Mammals
Pilot whale (n=7)
Harbor Porpoise
(n=5)
Beluga whale
(n=12)
Beluga whale
(n=12)
Northern fur seal
(n=2)
Ringed seal (n=4)
Beluga whale
(neonate) (n=l)
North Atlantic
North Atlantic
Arctic
Cook Inlet
North Pacific
Arctic
St. Lawrence estuary near
Quebec
Since
1987
Since
1987
Since
1987
Since
1987
Since
1987
Since
1987
1991
3847 (942-7118) ng/g (f.w.)
3260 (1880-4900) ng/g (f.w.)
1415 (142-2230) ng/g (f.w.)
624 (65.9-1630) ng/g (f.w.)
1190 (1050-1330) ng/g (f.w.)
198 (27-350) ng/g (f.w.)
689 ng/g (brain); 2289 ng/g
(kidney); 3370 ng/g (liver); 2106
ng/g (fat)
Becker et al.
(1997)
Becker et al.
(1997)
Becker et al.
(1997)
Becker et al.
(1997)
Becker et al.
(1997)
Becker et al.
(1997)
Gauthier et al.
(1998)
* Source: ATSDR (2002)
f.w. = fat weight basis; n = number; w.w. = wet weight basis
5.1.3 Intake of DDE from Food
Based on the U.S. FDA Adult Total Diet Study for October 1979-September 1980, the
daily intake of DDE was 0.004 |ig/kg in 1979 and 0.003 |ig/kg body weight/day in 1980
(Gartrell et al., 1986a). This study used the diet of a 16- to 19-year-old male as a basis for the
adult intake. Similar studies reported a daily intake of DDE for infants of 0.034 |ig/kg body
weight/day, and for toddlers of 0.045 |ig/kg body weight/day for 1980 (Gartrell et al., 1986b).
The dietary DDE intakes for eight population groups determined from more recent U.S. FDA
Total Diet Studies (from 1984 to 1991) are summarized in Table 5-3 (Gunderson, 1995a,b).
Table 5-3 Mean Daily Intake of DDE from Food Per Unit Body Weight (ug/kg body
weight/day) for Various Age Groups in the United States*
Analyte
6-11 mo
2yr
14-16 yr
F
14-16 yr
M
25-30 yr
F
25-30 yr
M
60-65 yr
F
60-65 yr
M
1984-1986
o,p'-DDE
p,p'-DDE
0.0002
0.0468
0.0001
0.0484
O.OOOl
0.01949
O.OOOl
0.0207
O.OOOl
0.0123
O.OOOl
0.0150
0.0001
0.0105
O.OOOl
0.0119
DDE —January, 2008
5-5
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Analyte
6-11 mo
2yr
14-16 yr
F
14-16 yr
M
25-30 yr
F
25-30 yr
M
60-65 yr
F
60-65 yr
M
1986-1991
o,p'-DDE
p,p'-VVE
O.0001
0.0441
<0.0001
0.0420
O.0001
0.0130
O.OOOl
0.0151
O.0001
0.0099
O.0001
0.0119
O.0001
0.0082
O.0001
0.0096
*Source: Gunderson (1995a,b)
F = female; M = male; mo = month; yr = year
5.2 Exposure from Air
Before 1972, when DDT was banned in the United States, it was used extensively as a
pesticide and large amounts of the chemical were released to air during agricultural or vector
control applications. Release of the chemical also could have resulted from production,
transport, and disposal operations. Due to the U.S. ban on DDT, current release in the United
States should be negligible (ATSDR, 2002).
However, an analysis of peat cores, as well as rain and snow samples from bogs or peat
lands across the mid latitudes of North America, indicated that DDT is still present in the
atmosphere (Rapaport et al., 1985). Because these areas receive all of their pollutant input from
the atmosphere, these data suggest that DDT still is being released to the atmosphere either from
its current production and use in other countries and transport to the U.S., or from the
volatilization of residues resulting from previous use in the US (ATSDR, 2002).
The estimated half-life of vapor-phase DDE is 17 hours; however, this estimate does not
necessarily reflect the lifetime of the compound in air. DDE can be adsorbed on particulate
matter, which slows photooxidation and results in long-range transport. Such long-range
transport has been demonstrated for DDT and several of its metabolites (Bard, 1999; Bidleman,
et al.,1992; Goldberg, 1975; Ottar, 1981; Wania and MacKay, 1993).
DDT, DDE, or DDD also has been detected in air samples from 32 of the 441 NPL
hazardous waste sites where those compounds were detected in some environmental media
(HazDat, 2002).
5.2.1 Concentration of DDE in Air
DDE is ubiquitous in the environment due to its persistence. In remote areas away from
agriculture DDE can be detected in air samples, over 20 years after the date it was banned in the
United States. Hawthorne et al. (1996) measured the DDE concentration in samples from two
rural sites between 1992 and 1994 in North Dakota at least 0.4 km from agricultural application
of pesticides. They found the chemical concentration ranged from 6 to 200 pg/m3 (Table 5-4);
the detection limit was 0.3 pg/m3. DDE concentrations varied by month, and levels detected in
1994 were greater than those from 1993.
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5-6
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Table 5-4 DDE Concentrations in North Dakota Air Sampling Collected in 1993 and
1994
1993
Temperature (°C)
DDE (pg/m3)
May 12
17
6.2
June 22
22
18
Augl
14
21
Augl7
23
20
Sept 21
12
8.1
Oct26
1
8.9
1994
Temperature (°C)
DDE (pg/m3)
May 18
26
9.7
July 18
23
120
Augl
24
49
Augll
22
41
AuglS
22
50
SeptS
22
200
Sept 22
8
8.8
Source: Hawthorne et al. (1996)
Another study conducted by Harner (2001) in Alabama, a region where DDT was used
heavily, DDE median air concentration in 1996 was 10 pg/m3 with a range of 1-92 pg/m3.
An air monitoring study in the Arctic region (Fellin et al., 1996) detected mean DDE
concentrations of 1.2 pg/m3 for the warm months (May-September, 1992) where the
temperatures ranged from -10 to 15°C (mean -3°C) and 2.9 pg/m3 cold months (October-April,
1992), with temperature ranges from -45 to -5°C (mean -25°C), respectively. These findings
seem counter to other temperature effects on DDE in the atmosphere. The increase of DDE and
other POPs in the atmosphere during the colder months in the Arctic are attributed to
anthropogenic contaminants from temperate latitudes being transported to the Arctic via
semi stationary high pressure cells that preside over Eurasia and North America. The advection
of these anthropogenic contaminants into the Arctic has been described as "Arctic haze" (Barrie,
1986).
DDE's volatilization is affected by temperature; it is expected that increasing
temperatures and moisture lead to higher ambient concentrations. This conclusion is supported
by data reported by Iwata et al. (1993) who collected and analyzed 71 samples of air over several
oceans (18 sampling locations) from April 1989 to August 1990. The author notes that at a
number of sampled locations ZJDDT concentrations were below the detection limit during most
of the winter months, while they were the highest during the summer months. An analysis of the
Great Lakes region between November 1990 and October 1991 demonstrated that the highest
monthly concentrations of DDE in Michigan cities were 63 pg/m3 in Saginaw (August), 119
pg/m3 in Sault Ste. Marie (May), and 92 pg/m3 in Traverse City (July; Monosmith and
Hermanson, 1996). The mean concentrations of DDE ranged between 0.3-180 pg/m3 and the
maximum was 180 pg/m3.
An analysis of the Monosmith and Hermanson (1996) data suggests a correlation
between higher DDT and DDE levels and air mass movement from the south, perhaps from areas
where DDT was still used (i.e., Central America or Mexico). As a comparison, the DDE level in
1989 was 15 pg/m3 over Green Bay, Wisconsin and 59 pg/m3 over the four lower Great Lakes
DDE —January, 2008
5-7
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(McConnell et al., 1998). In this study, however, an analysis of air masses indicated that the
atmospheric sources were local or regional volatilization, not long-range transport.
Atmospheric levels of pesticides in the United States were measured during a time of
high DDT usage. Stanley et al. (1971) took samples from 9 locations representing both urban
and agricultural areas. The group reported maximum o,//-DDT andp,p'-DDE levels in the range
between 2.4 and 131 ng/m3, with the highest levels found in the agricultural areas of the South.
Periodical monitoring of some agricultural areas in which DDT was extensively used has
provided data on the declining levels of the pesticide and its metabolites. DDE concentrations in
air over agricultural lands in Mississippi have declined from 2.6-7.1 ng/m3 as measured in 1967,
to 0.13-1.1 ng/m3, as measured in 1995 (Coupe et al., 2000). However, the authors point out that
those results also are evidence for the persistence of the DDT degradation products.
DDE levels were measured in air and rain in Portland, Oregon, in 1984 (Ligocki et al.,
1985). The authors did not detect DDE in rain samples; however, it was detected in five of
seven samples from the gas phase associated with the rainfall. Levels ranged from not detectable
to 420 pg/m3.
In 1986-1988, EPA assessed non-occupational exposure to pesticides (NOPES) for the
residents of two sites, Jacksonville, Florida and Springfield/Chicopee, Massachusetts (Whitmore
et al., 1994). The authors report that indoor DDE levels in air were higher than outdoor levels,
and that DDE was detected most frequently in the spring in Jacksonville (14%) and in the winter
in Springfield/Chicopee (20%). The estimated mean air DDE concentrations were < 1.0 ng/m3.
5.2.2 Intake of DDE from Air
DDT and its metabolites are fairly ubiquitous in the atmosphere; however, the
concentrations are so low that exposure via inhalation is generally negligible when compared to
other routes of exposure.
5.3 Exposure from Soil
In the U.S., DDT was primarily used as an agricultural pesticide. Consequently, large
amounts of DDT were released to agricultural soils; estimates include 27 million pounds in 1966
and 14 million pounds in 1971, shortly before it was banned (Gianessi et al., 1992). In addition,
direct or indirect releases during manufacturing, formulation, storage, or disposal of the chemical
have been sources of DDT in soils (ATSDR, 2002).
DDT and its metabolites are largely insoluble in water, so adsorb strongly to sediments in
aquatic systems or soils. The strong binding of DDE to soil particles suggests that it is more
likely to remain in the surface layers of soil, with little leaching into the lower soil layers and
groundwater (ATSDR, 2002). DDT, DDE, and ODD may revolatilize from soil to the air
(especially from moist soils) and then redeposited elsewhere by wet or dry deposition. This
DDE — January, 2008 5-8
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process of revolatilization and redistribution of DDT and its metabolites leads to the occurrence
of these compounds in areas where DDT never was used. This process is referred to as "global
distillation" and results in redistribution of the compounds from warm source areas to cold polar
regions.
As discussed in Chapter 3, the half-life of DDE is shorter in tropical soils than temperate
soils. DDT, DDE, or DDD has been detected in soil samples at 305 of the 441 NPL hazardous
waste sites where they were detected in the environment (HazDat 2002).
5.3.1 Concentration of DDE in Soil
Sediment
Monitoring - National
As discussed in Section 4 (above), the NAWQA program investigated the occurrence of
p,p'-DDE and 0.//-DDE, in bed sediments (Tables 4-3, 4-5). Sampling was conducted at 1310
sites from 1992 to 2001. The highest concentration of/\p'-DDE, 440 |ig/kg dry weight, was
found in a mixed land use setting; the highest concentration of 0,/>'-DDE, 26.7 |ig/kg dry weight,
was found in an urban setting (Nowell, 2003).
As part of the Environmental Monitoring and Trends Program (EMAP) 168 sites were
sampled along the southeastern coast of the United States in 1994-1995. The maximum
concentration ofp,p'-DDE in sediment was 34.2 |ig/kg, while the median concentration was
below the detection limit (Hyland et al., 1998). The National Surface Water Monitoring
Program monitored DDE in surface water and sediment in 1976-1980. The percent occurrence
(and maximum concentrations) of the DDE isomers in sediments was p,p' -DDE: 22.7% (163.0
jig/kg), and o,//-DDE: 0.5% (1.3 jig/kg) (Carey and Kutz, 1985).
Mixed use areas - Bays
Between February 1990 and March 1993, Gills et al. (1995) assayed 246 surface and
buried sediment samples from Newark Bay, New Jersey, and its major tributaries for the
presence and levels of DDT, DDE, and DDD. The mean DDE concentrations in surface
sediments ranged from 5 to 111 |ig/kg. In addition, the group calculated profiles of mean
sediment concentrations by the decade from 1940 to the present. DDE was rarely detected in
sediments deposited prior to 1940. Overall, its concentration in sediments deposited after 1970
(when DDT was banned in the United States) was lower (less than 500 |ig/kg) than in sediments
deposited between 1940 and 1960, the peak time period for production and usage of DDT-
containing insecticides in the United States. Although at lower concentrations, DDE occurred
more frequently in post-1970 sediments than in layers from the 1950s and 1960s (Gills, 1995).
In 1993, Pereira et al. (1996b) collected and analyzed bed sediments from sixteen sites
along the Lauritzen Canal and Richmond Harbor in the San Francisco area. Reported p,p'-DDE
values ranging from 1.9 to 860 ng/g dry weight of soil and o,//-DDE values ranging from 0.3 to
120 ng/g dry weight. DDE was detected at virtually every site tested (Pereira et al., 1996b).
Between 1975 and 1980, the USGS and EPA cooperatively monitored levels of pesticides in
DDE — January, 2008 5-9
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water and sediment at Pesticide Monitoring Network stations. Detectable levels of DDE were
recorded for 42 of the 171 stations (approximately 900 samples) monitored (Gilliom, 1984).
Between 1980 and 1983, EPA's STORET database listed analytical results for approximately
1,100 samples of sediments for DDE and reported median levels of 0.1 ng/kg dry weight
(Staples et al., 1985).
Mixed use areas - Watershed
The concentration of DDE in bed sediment from the San Joaquin River and its tributaries
in California (7 sites) in 1992 was 1.4-115 ng/L, respectively (Pereira et al., 1996a). Soil and
sediment samples (n = 28) were collected in 1987 from the Upper Steele Bayou Watershed in
west-central Mississippi at two depths (2.54-7.62 cm and 25.40-30.48 cm; Ford and Hill, 1991).
DDE was detected in the upper soil layers at an occurrence of 93%, with a mean concentration of
100 |ig/kg (range, non-detectable - 660 |ig/kg); in deeper soil layers the occurrence was 79%,
with a mean concentration of 40 |ig/kg (range, non-detectable - 560 |ig/kg).
Mixed use areas - Industrial (DDT production)
Some of the highest levels of DDE isomers in surface sediment (0-2 cm) were measured
in the Palos Verdes Shelf, near Los Angeles, where waste from a large DDT manufacturer was
discharged via a sewer outfall. The five sites in the area had o,p'-DDE and/?,//-DDE levels that
were between 6-45 mg/kg and 10-327 mg/kg, respectively (Venkatesan et al., 1996).
Soil
Monitoring - National
The U.S. National Soils Monitoring Program monitored the overall pattern of DDT
residues in soil, taking approximately 1500 samples each year. The 1970 results for/?,//-DDE
were 31% occurrence, with a mean concentration of 50 |ig/kg (range 10-6820 jig/kg). Those for
o,//-DDE were 3% occurrence, with a mean concentration of <10 |ig/kg (range non-detectable-
510 |ig/kg) (Crockett et al., 1974). The mean 2DDT level in five U.S. cities ranged from 120 to
560 |ig/kg in 1971, and generally urban areas had higher pesticide levels than did nearby
agricultural areas. The exceptions to this observation were some southern cities that were near
areas where the agricultural use of pesticides traditionally was heavy (Carey et al., 1979). The
Aberdeen pesticides dumpsite, Moore Country, North Carolina, had some of the highest reported
surficial soil levels of DDE that were between 8.5 |ig/kg and 2335 |ig/kg (Vine, 2000).
Agricultural
Samples of 38 soils collected between 1995-1996 from the corn belt in the mid-central
United States where DDT was heavily used, showed a geometric mean concentration for
p,p'-DDE of 3.75 |ig/kg (Aigner et al.,1998). The DDT/DDE ratio was determined in 21 of the
samples and ranged from 0.5 to 6.6. Forest soils in Maine, which had been sprayed with DDT in
the past, had 2DDT levels between 270 and 1898 |ig/kg, much higher than the maximum
concentration of 11 |ig/kg in unsprayed locations.
Soil samples taken near Dell City, Texas, in 1980 contained an average of 0.46 mg/kg
(dry weight) of DDE (Hitch and Day, 1992). DDT use was extensive in agricultural areas in
DDE — January, 2008 5-10
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Arizona for 18 years until a statewide moratorium on use in January 1969, after which the
pesticide and its residues were monitored. Three years later, residues in agricultural soils had
decreased by 23%, and the ratio of DDE to DDT was increasing, indicating a transformation of
DDT to DDE (Ware et al.,1978). Similar results were reported by Buck et al. (1983) from
monitoring these same sites over 12 years following the ban on DDT use. During that period,
combined DDT and DDE residues in agricultural soils had fallen from 1.2 to 0.39 mg/kg, while
those in surrounding desert soil had fallen from 0.40 to 0.09 mg/kg. Yet another study (Aigner et
al., 1998) reports a mean DDE concentration of 3.75 ng/g soil as measured in 38 agricultural and
2 garden samples from Pennsylvania, Ohio, Indiana, and Illinois.
A study measured organochlorine pesticide concentrations in elementary school yards
along the Texas-Mexico border. It concluded that DDE was the most frequently detected
contaminant, occurring at 69% of the 13 sites examined. Relatively higher concentrations were
observed in agricultural areas along the southern border (50-60 ppb in soils from Harlingen,
McAllen, Palmview, and San Benito) than in other soils. In contrast, the DDE range from
reference soil samples collected from eight national parks was between 4 and 9 ppb (Miersma,
2003). Historical data obtained during the time of DDT use indicate that its concentration in soil
ranged from 0.01 to 53 ppm (Miersma, 2003).
5.3.2 Intake of DDE from Soil
Specific studies were not found in the literature reviewed concerning the intake of DDE
from soil. An estimate of the intake ofp,p'-DDE from soil can be derived from the data
produced by the U.S. National Monitoring Program, which reported concentrations that ranged
between 0.01 to 6.82 mg/kg of soil (Crockett et al., 1974). Intake can be estimated assuming that
a 70-kg adult ingests 50 mg of soil daily and a 10-kg child ingests 100 mg (U.S. EPA, 1997a).
Based on these data and assumptions, the intake of p,p'-DDE from soil for adults ranges between
0.000007 and 0.005 |ig/kg-day, and for children it is between 0.0001 and 0.07 |ig/kg-day.
5.4 Other Residential Exposures
Monitoring in older homes reveals that carpeting in these homes may have high levels of
DDE (Lewis et al., 1994). In one house built in 1930, the carpeting, which was believed to be at
least 25 years old, contained up to 10.8 |ig/m2 or 5.7 |ig/g of 2DDT (/\p'-DDT, ODD, and
DDE).
Chlorinated pesticide residues also have been detected in cigarettes. Djordjevic et al.
(1995) measured their content in U.S. and foreign cigarettes manufactured between the 1960s
and the 1990s. Since 1970, the concentration of DDT analogs has decreased by more than 98%.
The levels of o,//-DDD detected in cigarettes manufactured between 1961-1979 were 396-7150
ng/g; by 1983-1994 levels were reduced to levels that were undetectable up to 19.0 ng/g. Levels
ofp,p'-DDE in cigarettes were 58-959 ng/g between 1961-1979; they also were reduced by
1983-1994 to 6.6-15.8 ng/g. The same author also measured the transfer rate of DDE from
tobacco into mainstream smoke to be 27%.
DDE — January, 2008 5-11
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ATSDR (HazDat, 2002) reports that DDT, DDE, or ODD have been identified in at least
441 of the 1613 hazardous waste sites proposed for inclusion on the EPA National Priorities List
(NPL). The number of sites evaluated for those compounds was not specified, but/\p'-DDT,
p,p'-DDE, and o,p'-DDE were detected at 326, 219, and 4 sites respectively (U.S. EPA, 1994c).
5.5 Occupational (Workplace) Exposures
DDT is not produced or used in the United States; therefore little or no occupational
exposure to DDT or its degradates, including DDE, is expected via manufacturing or application.
In 1971, the estimated respiratory exposure potential for formulating plant workers was 14.1
mg/person/hour (Wolfe and Armstrong, 1971)
5.6 Summary
Due to the persistence of DDE in the environment, exposure is likely to occur through
media in addition to water, including exposures through food, air, soil, and residential contact.
The general population likely has exposures primarily through food. Because of the persistence
of DDT and DDE, it is anticipated that low levels of residues will be present in commodities for
decades. In fact, depending on the continued use of DDT and export of commodities that may
contain DDT by other countries, levels in the diet may even increase (Coulston, 1985). In
domestic commodities, dietary exposure of consumers may result from residues bioaccumulated
in some food items, including fish.
DDT and its metabolites are transported globally in the atmosphere but are present in
such low concentrations that exposure via inhalation is negligible. Soil levels of DDE have been
gradually declining; an estimate for adultp,p'-DDE intake from soil ranges between 0.000007
and 0.005 |ig/kg-day, and for children it is between 0.0001 and 0.07 |ig/kg-day. Other exposures
may be residential in origin, including residues found in carpeting and cigarette smoke.
DDE — January, 2008 5-12
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6.0 TOXICOKINETICS
6.1 Absorption
Oral Exposure
No studies were located that quantify the rate or extent of absorption of DDE in either
humans or animals after oral exposure. Gastrointestinal absorption, however, can be inferred by
measurements in serum and adipose tissues, the presence of metabolites in the urine, and toxic
effects observed (ATSDR, 2002; Fawcett et al., 1987; Gold and Brunk, 1982, 1983, 1984; Hayes
et al., 1971; Morgan and Roan, 1971, 1974). DDE is preferentially absorbed by the intestinal
lymphatic system; however, some absorption into the portal blood also occurs (ATSDR, 2002).
DDT has been shown to be absorbed 1.5-10 times more effectively in laboratory animals when
administered in digestible oils compared to administration in nonabsorbable solvents (Hayes,
1982); this is likely also true for DDE. Sieber (1976) demonstrated vehicle-related differences in
the uptake of DDE into the lymphatic system. Uptake via the chylomicrons was 30% when the
DDE was in an aqueous vehicle and 60% when in corn oil.
Dermal Exposure
No studies that quantify the rate or extent of dermal absorption of DDE in either humans
or animals were located. However, DDT is poorly absorbed through the skin even in solution
(WHO, 1979).
Inhalation Exposure
No studies were located that quantify the rate or extent of absorption of DDE in either
humans or animals after inhalation exposure.
6.2 Distribution
Once absorbed, the distribution of DDE is similar regardless of route of exposure. Both
the blood and lymph deliver DDE to all body tissues with storage in these tissues generally
proportional to the lipid content of the tissue (Morgan and Roan, 1971). DDE was demonstrated
to be stored at a greater rate than either DDT, or DDD, a companion DDT metabolite (Morgan
and Roan, 1971).
Although DDE transport from the intestines involves distribution to both portal
circulation and the lymphatic system, the proportions vary with the nature of the exposure
vehicle (oil or water; Sieber, 1976). Systemically less than 18% of DDE is carried in human
erythrocytes. Gomez-Catalan et al. (1991) found 82.3% of DDE in human blood associated with
plasma and 17.7% with cells. In plasma, 12% of the DDE recovered was associated with low
density lipoproteins (LDL), 9% with very low density lipoproteins (VLDL), and 6% with high
density lipoproteins (HDL). Noren et al. (1999) found that DDE and its methylsulfonyl
metabolite, 3-MeSO2-DDE, were primarily (80%) in the albumin fraction of human blood.
Gomez-Catalan et al. (1991) estimated that about 60% of the DDE in plasma was associated with
proteins and 9-15% with lipoproteins.
DDE — January, 2008 6-1
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Because of the high fat content in human breast milk, DDE is selectively partitions into
it. DDE was detected in all human breast milk samples collected and tested from 1969 to 1970 in
a U.S. national human milk study. The mean concentration of DDE in human breast milk was
1.9 ppm (lipid-basis; Takei et al., 1983). Using a model, the body burden of DDE in infants
exposed via mother's milk was estimated to increase rapidly at the start of lactation, but decrease
after approximately 5-6 months, even with continued nursing under all the exposure scenarios
evaluated. Maximum mean body burden of DDE was estimated to be 70 |ig/kg with the level
reduced to <10 |ig/kg by 24 months regardless of the duration of breast-feeding (LaKind et al.,
2000).
DDE is distributed to the fetus during pregnancy. In one study, lipid-adjusted mean
concentrations were similar in maternal blood and cord blood (4.4 ppm maternal; 4.7 ppm cord
blood; Waliszewki et al., 2000). However, O'Leary et al. (1970b) and Schvartsman et al. (1974)
found that cord blood levels were lower than the corresponding maternal blood. Sala et al.
(2001) reported a geometric mean of 2.2 ppb (not lipid-adjusted) for DDE at delivery in maternal
blood of 72 Spanish women, who lived in the vicinity of an organochlorine-compound plant
compared to 0.83 ppb in cord blood (n=69).
DDE exposure of pregnant rats demonstrated that very little DDE crossed the placenta.
Pregnant rats were administered 10 or 100 mg/kg/day on gestational days (gd) 14-18. In dams
sacrificed on gestational day 15 or 17, the concentration of DDE in the placenta was about 3-fold
higher than in fetal tissues. Ten-day-old pups exposed only in utero to 10 mg/kg/day had no
detectable levels of DDE in blood, liver, or brain, while the pups exposure to 100 mg/kg/day had
measurable DDE in their livers, but not in their blood or brain (You et al., 1999c).
In the study by You et al. (1999c) some pups were exposed to DDE through combined
gestational and lactation exposure or during lactation only. There were no statistically-
significant differences between the two groups in the levels of DDE detected in the liver, brain,
and blood. By 78 days of age, all three exposure groups had DDE detectable only in fat
deposits. Rats exposed only in utero had at least 2-fold less DDE in their fat than the other two
exposure groups. DDE levels in the tissues and plasma of dams at the end of nursing were
approximately a third those observed at the end of gestation, suggesting mobilization of DDE
from storage sites during lactation.
Distribution kinetics of intravenously administered DDE (5 mg/kg) demonstrate a
redistribution from blood to liver and muscle, to skin, and ultimately to adipose tissue with the
entire process appearing to take about 1 day. Peak concentrations of DDE were observed before
1 hour in the liver and muscle, at 3 hours in the skin, and between 1 and 4 days in adipose tissue.
The tissue/blood concentration ratio 4 to 14 days after injection was 6 for liver and muscle, 35
for skin, and 400 for adipose tissue (Miihlebach et al., 1991).
DDE — January, 2008 6-2
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6.3 Metabolism
Overview of Metabolic Pathways
The metabolism of DDE has been studied in humans and a variety of other mammalian
species. The metabolism in rats, mice, and hamsters is similar to that in humans; however, not
all of the intermediary metabolites have been identified in both humans and animals. DDE is
metabolized to easily excretable phenols, as well as to m-methylsulfone-/\p'-DDE. DDE has
been demonstrated to induce hepatic cytochrome P-450 (mainly CYP2B).
Studies of DDE Metabolism in Humans
Studies by Roan et al. (1971) and Morgan and Roan (1971) indicate that in humans little
if any DDE is converted to bis(p-chlorophenyl)acetic acid (DDA). Part of stored DDE is
excreted unchanged in humans; however, because its elimination is promoted by the induction of
microsomal enzymes, it appears that it undergoes metabolism to include conjugation.
Animal Studies of DDE Metabolism
It has been proposed that in mammals, including humans, that DDT is initially
metabolized to DDE (Mattson et al., 1953; Pearce et al., 1952) and l,l-dichloro-2,2-bis(p-
chlorophenyl)ethane (DDD) (Klein et al., 1964) in the liver (Figure 6-1). Further metabolism of
DDE is apparently slow with DDE retained in adipose tissues (Hayes et al., 1971; Morgan and
Roan, 1971). In cases where the initial exposure is to DDE, especially when it is dissolved in
corn oil, a considerable portion of the DDE could partition to adipose tissues before reaching the
liver for further metabolism since chylomicrons are discharged directly from the lymphatic
system to systemic circulation.
In rats, DDE is slowly converted in the liver to l-chloro-2,2-bis(p-chlorophenyl)ethene
(DDMU). DDMU is converted to l,l-bis(p-chlorophenyl)ethene (DDNU), which is metabolized
in the kidney to 2,2-bis(p-chlorophenyl)ethanol (DDOH), followed by 2,2-bis(p-
chlorophenyl)ethanal (DDCHO) (Suggs et al., 1970), and finally oxidized to DDA (Peterson and
Robison, 1964). DDA can then go through conjugation with glycine, bile acid conjugates,
serine, aspartic acid, or glucuronic acid prior to elimination (Gingell, 1975; Pinto et al., 1965;
Reif and Sinsheimer, 1975). It also has been suggested that conversion to DDA may proceed by
way of an acid chloride intermediate (DDA-C1) in Wistar rats, mice, and hamsters which would
be hydrolyzed to form the acid. (Fawcett et al., 1987; Gold and Brunk, 1982, 1983, 1984).
An alternate metabolic option for DDMU is conversion to a DDMU expoxide. It is
hypothesized that the epoxide could generate an intermediate with an electrophylic carbon
(Datta, 1970; Datta and Nelson, 1970) capable of adding to and modifying DNA or other
macromolecules involved in cell cycle control and/or signal transduction.
Methylsulfonyl metabolites also have been identified. Two such metabolites (3- and 2-
methylsulfonyl-DDE) have been isolated in the blubber of seals from the Baltic (Jensen and
Jansson, 1976), as well as in several other species (Bergman et al., 1994) including humans
(Weistrand and Noren, 1997). Figure 6-2 demonstrates metabolism of DDE into its
methyl sulfonyl metabolites. DDE in the presence of phase I enzymes produces arene oxide,
DDE — January, 2008 6-3
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which then undergoes conjunction with glutathione. Once excreted in the bile into the large
intestines, it is cleaved by microbial C-S lyase (Bakke et al., 1982; Preseton, et al., 1984) into a
thiol. The thiol is then methylated, reabsorbed, and the sulfur is further oxidized to
methylsulfones, which are distributed by the blood (Haraguchi, et al., 1989).
6.4 Excretion
There are limited data on human excretion of DDE. DDE was measured in the bile of 5
male human subjects (Paschal et al., 1974). DDE is eliminated in human breast milk most likely
due to its high lipid concentration. Lactational transfer is not an excretory pathway, but does
result in removal of DDE from the mother and transfer to the offspring or into the food supply.
DDE has been detected in the urine of mice and hamsters following both acute and
chronic exposure to DDT (Gingell, 1976; Gold and Brunk, 1983). Intravenously administered
DDE (5 mg/kg) was mainly excreted in the feces (34%) of rats within a 14-day period following
a single dose (Miihlebach et al., 1991); 1% was excreted in the urine. Although 10% of the DDE
excreted in the feces was unchanged, no unchanged DDE was detected in the urine and no
hexane-extractable lipophilic metabolites were detected in the feces. The total body burden half-
life for DDE in rats was estimated to be 120 days (Miihlebach et al., 1991) and is longer than
that for DDT (ATSDR, 2002).
DDE — January, 2008 6-4
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Figure 6-1 Proposed metabolic pathway for DDT, including further metabolism
of DDE (adapted from Peterson and Robinson, 1964 and presented in
ATSDR, 2002)
R2CHCC13
DDT
LIVER
LIVER
R,-C CrlCl
O
\
DDMU epoxide
electrophile
R2C=CC12
DDE
SLOW
LIVER
LIVER
R,CHCHC12
* ODD
FASTER
LIVER
R2OCHCI
DDMU
LIVER
R2CHCH2C
X PPMS
-MCI/ KIDNEY &
^f LIVKR (some)
R,C=CH2
DDNU
I
KIDNKY
R2CH-CH2OH
DDQH
1
-2H KIDNEY
R3CH-CHO
DDCHO
R2CH-COOH
PDA
* Adapted from Peterson and Robinson 1964
CONJUGATION
DDE —January, 2008
6-5
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Figure 6-2 Proposed metabolic pathway for the conversion of/>,/>'-
DDE to its methylsulfone derivatives (Bergman et al.,
1994; Letcher et al., 1998; Weistrand and Noren, 1997)
phase I
enzymes
P450Cyp-2B
phase I
enzymes,
glutathione
bile
flora
(C-S lyase)
methylation of
thiol,
oxidation of thiol
S-glutathione
SO2CH3
DDE —January, 2008
6-6
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7.0 HAZARD IDENTIFICATION
7.1 Human Effects
Human exposure to DDE can occur via human metabolism of DDT, the parent pesticide
of DDE, or via direct exposure to DDE that has been degraded from DDT in the environment
(e.g., through fish consumption). DDE is stored in fat leading to long-term internal exposure.
There are commonalities and differences in the health effects associated with DDT and DDE
based on experimental studies of both compounds in animal species. In cases where the initial
exposure was only to DDT, it is difficult to determine which effects are due to the DDT, DDE or
other metabolites.
7.1.1 Short-Term Studies and Case Reports
General Population
Intentional and Accidental Acute Ingestion
No studies of acute intentional or accidental ingestion of DDE in humans were found.
However, data on DDT can be useful in hazard identification for DDE. The only documented
case of human fatality following DDT ingestion occurred when a 1-year-old child ingested one
ounce of 5% DDT in kerosene (Hill and Robinson, 1945). In other reports, doses as high as 285
mg DDT/kg body weight have been accidentally ingested without fatality (Garrett, 1947).
Oral exposure of doses up to approximately 22 mg DDT/kg in an oil solution caused
neurological effects including prickling sensation of the area in and around the mouth,
disturbance of sensitivity to the lower part of the face, uncertain gait, malaise, cold moist skin,
hypersensitivity to contact, disturbance of equilibrium, dizziness, confusion, tremors, headache,
fatigue, and severe vomiting within 10 hours of exposure. By 24 hours after exposure, most
symptoms were no longer evident (Velbinger, 1947a,b). Perspiration, headache, and nausea
were reported following a single oral exposure to 6 mg DDT/kg (Hayes, 1982). Convulsions
have been reported in humans exposed to 16mg DDT/kg or higher (Hsieh, 1954). Similar
symptoms have been reported after accidental or uncontrolled intentional ingestion of DDT
(ATSDR, 2002).
To examine immunological effects, three volunteers were administered 5 mg DDT/day
(0.07 mg/kg) for 20 days, and then challenged with an injection of Salmonella typhimurium
vaccine. Volunteers receiving DDT had significantly higher serum agglutinin liters than
volunteers receiving vaccine alone; immunoglobulin levels were unchanged. There were no
apparent side effects associated with the DDT exposure (Shiplov et al., 1972).
7.1.2 Long-Term and Epidemiological Studies
No longer-term intentional dosing studies in humans were identified for DDE; however, a
12-18 month study of DDT administration to humans was identified (Hayes et al., 1956). In this
study doses of 3.5 or 35 mg DDT/day were administered to 51 male volunteers and no
DDE — January, 2008 7-1
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neurological (e.g., loss of coordination, tremor, etc.), cardiac (e.g., blood pressure, heart rate,
etc.), or body weight effects were reported.
A number of occupational and general population epidemiology studies have explored
the relationship of serum DDE levels and various physiological conditions or endpoints. The
majority of the studies examined cancer endpoints or those associated with reproduction and
development. Summaries of these studies and their findings follow.
General Populations
Noncancer Systemic Effects
In a case-control study of patients with chronic, debilitating fatigue lasting at least 6
months, the mean concentration of DDE in blood serum was significantly higher in cases (11.9
ppb; n=14) than in controls (5.2 ppb; n=23) (Dunstan et al., 1996). The 37 subjects were pooled
and then re-divided according to their serum DDE levels (high being >6 ppb and low being <6
ppb). There was greater variability in the width of the erythrocyte cells in the high DDE group
compared to the low DDE group; no differences were observed in other hematological
parameters. Variability in the width of the erythrocyte cells can be a sign of anemia; unresolved
anemia is one factor that is often associated with chronic fatigue.
Bohannon et al. (2000) hypothesized that high levels of DDE would be associated with
lower bone density in peri- and post-menopausal women than in premenopausal women due to
its (and DDT's) estrogen- and androgen-like properties. Women (50 black and 53 white) from
the Mount Sinai Medical Center Longitudinal Normative Bone Density Study (1984-1987), were
studied to examine the relationship between serum levels of DDE and bone mineral density. The
study found that black women had significantly higher serum DDE levels than white women
(mean 13.9 ppb vs. 8.4 ppb, respectively), but it did not find a correlation between DDE and
either bone density or the rate of bone loss in the lumbar spine over a 2-year period.
In a study of 68 sedentary Australian women (45-64 years of age), an association was
found between bone mineral (lumbar spine) and serum DDE levels >2 ppb. Only about 7.24%
of the variance, however, was attributed to DDE with a stronger correlation found between bone
mineral density and age (Beard et al., 2000). A study of 115 Swedish men aged 40-75 did not
find a correlation between serum DDE levels and bone mineral density (Glynn, 2000).
Immunotoxicity
Vine et al. (2001) measured immunological parameters in 302 individuals (residing near
a waste site) with a median of 2 ppb blood DDE (32 ppb blood DDE maximum). Parameters
included white blood cell counts, lymphocyte phenotypes, immunoglobulin, mitogen-induced
lymphoproliferative activity, and a skin test to evaluate delayed-type hypersensitivity. Although
20 organochlorines were targeted by the analysis, DDE was the only one detected. Results
indicated that subjects with higher blood DDE levels had lower mitogen-induced
lymphoproliferative activity (concanavalin A) and slightly increased total lymphocyte
immunoglobulin A levels. The other parameters studied were unaffected; the authors concluded
DDE — January, 2008 7-2
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that relatively low blood DDE levels were associated with changes in immune markers, but the
changes were of uncertain clinical significance.
Reproductive and Developmental Endpoints
Infants of 23 women with preterm deliveries had higher DDE blood levels (19-22.1 ppb)
than the full-term infants (4.9-6.1 ppb DDE) of 44 women. DDE was the only organochlorine
monitored (O'Leary et al., 1970a).
A dose-related trend (p<0.0001) was observed between DDE concentrations in maternal
blood and odds of pre-term birth among 2380 children born between 1959 and 1966
(Longnecker et al., 2001). The study cohort included 361 pre-term deliveries and 221 children
who were small-for-gestational-age. Serum samples were obtained from mothers during the
third trimester and stored at -20°C until 1997-1999 when the serum samples were analyzed. The
median serum DDE was 25 ppb (range of 3-178 ppb). Quadratic spline models showed that the
odds of pre-term birth began to increase at a blood DDE concentration of 10 ppb. Adjusted odds
for small-for-gestational-age births also increased, but the trend was less consistent (p=0.04); the
association of DDE with small-for-gestational-age births remained even with the exclusion of
pre-term births. The authors noted that because pre-term births are a major contributor to infant
mortality, inclusion of only children who survived may have lead to an underestimate of DDE's
effects.
DDE levels in maternal milk were not associated with birth weight, head circumference,
or neonatal jaundice; however, DDE at levels >4 ppm in maternal milk was associated with
decreased reflex reactions in infants (Rogan, 1986). The same authors found an inverse
relationship between the levels of DDE in maternal milk and lactation duration in women in the
United States and Mexico (Rogan, 1987). Gladen and Rogan (1995) examined the relationship
between DDE and lactation duration in Tlahualilo, Mexico and found an inverse relationship
between DDE and duration of lactation only applied to women with prior breast-feeding
experience and not in women breast feeding for the first time. A statistically-significant
decrease in median lactation duration was observed in women with > 12 ppm (lipid-adjusted; 3
months of lactation) compared to women with 0-2.5 ppm (7.5 months of lactation). The
observed effects on lactation may be due to the ability of DDE to disrupt normal hormonal
regulation of lactation (Guyton, 2000).
A prospective longitudinal study was performed to examine the relationship between
prenatal or lactational exposure to background levels of DDE (or poly chlorinated biphenyl,
PCB) and pubertal growth and development. Exposure was estimated for 856 children by
measuring DDE in maternal milk, maternal blood, cord blood, and placenta. These children had
normal birth weights and normal growth during their first year of life. Five hundred and ninety-
four (316 girls and 278 boys) were available at puberty for follow-up. The transplacental DDE
index ranged from 0.3 to 23.8 ppm (median 2.4 ppm) and the lactational DDE index ranged from
0.2 to 96.3 mg (median 6.2 mg). Transplacental exposure to DDE was associated with increased
height and weight of boys at puberty. Boys with the highest exposures to transplacental DDE (a
median index of 2.4 ppm) were 6.3 cm taller and 6.9 kg larger than boys from the lowest
DDE — January, 2008 7-3
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exposure (0-1 ppm). These effects were not observed in the girls. No effects on the age at which
pubertal stages were attained were observed (Gladen et al., 2000).
Karmaus et al. (2002) studied the association of serum DDE levels in children at 8 years
of age with weight and height through age 10. Medical records were used to obtain the height
and weight of the children from birth to 4 years of age. While there were 323 children with
weights and heights measured at birth, there were only 202 at 43-48 months. Measurements at 8,
9, and 10 years of age were performed by the investigators (the number of children at each age
was not provided in ATSDR, 2002). Blood was collected and analyzed only at 8 years of age.
Results indicate that, after controlling for relevant confounders, DDE was a significant predictor
of height in girls from 1.3 months until 8 years of age. Girls in the highest DDE quartile (0.44-
40.4 ppb) were 1.8 cm shorter than girls in the lowest quartile (0.08-0.2 ppb). At 10 years of
age, no association was found.
An evaluation of Inuit infants' in utero exposure to DDE (and other organochlorines)
revealed no association with immunological parameters, but there was a relative risk of 1.87
(95% CI, 1.07-3.26) for increased otitis media for 4- to7-month old infants in the highest tertile
of DDE exposure when compared to the lowest tertile (Dewailly, 2000). The organochlorine
concentration in breast milk was used as an index of the infants' exposure. The authors of this
study noted that there the organochlorine exposures in this population originate from the same
few food items, such that the concentrations of the various chemicals (DDE, hexachlorobenzene,
etc.) are correlated with each other.
Cancer Endpoints
A positive correlation was made between DDE and breast cancer in New York City
women attending a mammography clinic between 1985 and 1991. Serum DDE (not lipid-
corrected) was measured in archived blood samples of 58 women who were diagnosed with
breast cancer within 6 months of entering the study and in matched (by menopausal status, age,
number and dates of blood donations, and day of menstrual cycle at time of first blood drawing)
with cancer-free control women from the same cohort (n=171). Case patients had a significantly
(p=0.031) higher mean serum DDE (11.0 ppb) than the control subjects (7.7 ppb). The adjusted
odds ratio for breast cancer in the highest quintile of serum DDE was 3.68 (95% Confidence
Interval [CI]=1.01-13.50) using the lowest quintile as the referent group. A positive trend
(p=0.035) was observed in the odds ratio for breast cancer and increasing quintile, and a positive
trend (p=0.0037) was observed between the odds ratio and serum DDE when serum DDE was
elevated as a continuous variable using conditional multiple logistic regression. The odds ratios
were adjusted for confounders, such as first-degree family history of breast cancer, lifetime
months of lactation, age at first full-term pregnancy, age at menarche, history of benign breast
disease, history of tobacco and alcohol use, and race (Wolff et al., 1993).
Age-adjusted serum DDE levels were significantly related to breast cancer risk in 120
cases (3.84 ppm) compared to 126 controls (2.51 ppm) living in Mexico City. There was a
marginally significant (p=0.06) positive trend between DDE levels and the risk of breast cancer
(ORQ1.Q4=2.16; 95% CI=0.85-5.50) even after adjusting for age at menarche, duration of breast
feeding, quetelet index (as a measure of obesity), and menopausal status. The association was
DDE — January, 2008 7-4
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strongest in menopausal women, with no association found in premenopausal women. Serum
DDT levels did not have the same association. Major predictors of DDE levels included DDT
levels, age, duration of lactation, parity, and socioeconomic level (Romieu et al., 2000).
A few studies found an increase in DDE levels in malignant breast tissue in comparison
to adjacent "normal" breast tissue or to benign breast disease; however the difference in DDE
levels generally was not significant (Dewailly et al., 1994; Falck et al., 1992; Wasserman et al.,
1976). A case-control study in German women performed in 1993-1994 found a 62% (p=0.017)
higher level of DDE in malignant breast tissue of recently masectomized women (n=45)
compared to age-adjusted geometric mean DDE concentration in benign breast tissue in the
control group (n=20). A significant difference was not observed in the concentrations of DDT
(Guttes et al., 1998).
A hospital-based case-control study of Canadian woman found a weak association
between DDE in breast adipose tissue and breast cancer. The geometric mean of DDE in the 217
cases was 693 ppb compared to 596 ppb in the 213 benign controls. The increased risk was
observed when current hormone replacement therapy users were excluded. The odds ratio of
DDE was higher for the risk of estrogen receptor-negative breast cancers than for estrogen-
positive breast cancer (Aronson et al., 2000; Woolcott et al., 2001).
Although there are several studies associating DDE with breast cancer, there are also
many showing no association, even when the studies were performed by the same authors.
These studies adjusted for confounders, had matched controls, used current or past (up to 19
years prior) DDE levels, used serum or adipose DDE concentrations, and had sufficient numbers
of cases (Demers et al., 2000; Helzlsouer et al., 1999; H0yer et al., 1998; Hunter et al., 1997;
Laden et al., 2001a; Liljegren et al., 1998; Lopez-Carrillo et al., 1997; Matuo et al., 2000;
Mendonca et al., 1999; Moysich et al., 1998; Schecter et al., 1997; Unger et al., 1984; Ward et
al., 2000; Wolff et al., 2000a,b; Zheng et al., 1999, 2000).
Hunter et al. (1997) performed a prospective study on the health of 121,700 married
nurses in the United States. Cholesterol-adjusted plasma DDE was determined in blood samples
collected from 1989 to 1990. There was no significant difference in historical plasma DDE
(1989-1990) in women who developed breast cancer before June 1992 (n=236) and pair-wise
matched control women who did not develop breast cancer. This lack of association was
observed regardless of menopausal status, age, age at menarche, age at birth of first child,
number of children, and history of lactation. A follow-up report added 143 postmenopausal
cases and controls, and adjusted for plasma organochlorines and for triglycerides in addition to
cholesterol (Laden, 200 Ib). Median (lipid-adjusted) concentrations of DDE in cases and
controls were 0.768 and 0.817 ppm, respectively. Again, no association between breast cancer
and DDE was observed within strata of age, age at menarche, age at birth of first child, number
of children, history of benign breast disease, or family history of breast cancer. The multivariate
relative risk of breast cancer for women in the highest quintile of DDE (1.5-6.0 ppm) compared
to women in the lowest DDE quintile (0.007-0.43 ppm) was 0.82 (95% CI=0.49-1.37).
DDE — January, 2008 7-5
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Krieger et al. (1994) did not find an association between serum DDE levels and breast
cancer in 150 cases (50 each of Caucasian, African-American, and Asian women) compared to
150 matched controls (43.1 vs. 43.3 ppb; not lipid adjusted) in a prospective nested case-control
study. Blood was collected between 1964 and 1971 from women in northern California.
Controls did not develop breast cancer through the end of 1990. When the ethnic groups were
examined separately, however, high serum DDE concentrations were associated with breast
cancer incidence in Caucasian and African-American women (not significant), but not Asian
American women. Black women with breast cancer had an average of 5.7 ppb higher levels of
serum DDE than paired controls, but the difference was not statistically significant (95% CI on
the difference ranging from -3.3 to 14.8). It also should be noted that serum DDE concentrations
were higher in black and Asian women compared to white women.
A case-control study nested within a prospective cohort study was conducted in residents
of Washington County, Maryland, to examine the association between breast cancer and serum
DDE (both lipid-adjusted and unadjusted) from blood samples collected in 1974 (20,305
samples) or 1989 (25,080 samples). The cases consisted of 346 women diagnosed with breast
cancer by June 1994. Controls (n=346) were cancer-free women matched for age, race,
menopausal status, and date of blood donation. Lipid-adjusted DDE concentrations were higher
(nonsignificantly) in the controls than in the cases. Risk analyses conducted on quintiles (1974
samples) or tertiles (1989 samples) showed no association between DDE (lipid-adjusted) and
breast cancer (Helzlsouer et al., 1999).
Wolff et al. (2000a) investigated breast cancer risk associated with organochlorine
exposure in a hospital-based, case-control study using 175 cancer patients, 181 control patients
with benign breast disease, and 175 women in a second hospital control. Overall 57% of the
subjects were Caucasian, 21% hispanic, and 22% African-American. Several organochlorines
were measured in blood serum, which generally was obtained before surgery and no more than 2
months after surgery and measured for 4 tumor markers (estrogen, progesterone, p53, and erbB-
2). Results demonstrated that African-American woman had significantly higher lipid-adjusted
DDE levels (geometric mean of 1.0 ppm) than Hispanic (0.71 ppm) and Caucasian (0.48)
women. Organochlorine levels were not associated with breast cancer risk, including tumor
stage or tumor markers. Higher DDE levels were associated with increasing body mass index
(BMI), as well as, decreasing level of education, frequency of nulliparity, and frequency of
family history of breast cancer. The findings were attributed to historical patterns of exposure
and to metabolic differences in organochlorines related to BMI.
In a re-evaluation of the data from their 1993 report, the same investigators (Wolff et al.,
2000b) did not find the same results. In the re-evaluation, the authors were interested in
assessing the half-lives of DDE and PCBs; therefore, the study was restricted to cases with at
least three yearly blood donations after the first year of serum analysis. Lipid measurements
were available in 110 cases and 213 controls, and DDE half-life was calculated for 84 cases and
196 controls. Odds ratios were calculated with conditional logistic regression analysis using
DDE (as well as PCB) as both quartile or tertiles and continuous variables. DDE levels were
similar in the cases (geometric mean 977 ppb, lipid-adjusted) and controls (1097 ppb) regardless
of whether the concentration was lipid-adjusted or not. Differences were not noted when
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estrogen receptor status of tumors, age at menarche, number of full-term pregnancies, age at first
full-term pregnancy, first degree family history of breast cancer, months of lactation, and height
were considered. The half-life of DDE, but not serum level, was correlated with BMI.
A nested case-control study of 240 breast cancer patients and 477 control subjects from a
prospective heart study of 7712 women in Denmark found that odds ratios for breast cancer were
not increased in the upper quartiles of lipid-adjusted serum DDE (H0yer, 1998). Breast cancer
diagnosis occurred up to 17 years following the collection of blood samples. The analysis
considered the following confounders: weight, height, number of full term pregnancies, alcohol
consumption, smoking, physical activity, menopausal status, household income, marital status,
and education.
Ward et al. (2000) also did not find an association between breast cancer and serum DDE
using 150 breast cancer cases and 150 controls from Norway. Blood was collected 2 to 18.2
years (mean of 8.8 years) prior to diagnosis. Mean lipid-adjusted DDE concentrations were
1230 and 1260 ppb in cases and controls, respectively (corresponding DDT concentrations were
120 and 138 ppb, respectively). When the data was stratified by age at diagnosis, the interval
between blood collection and diagnosis, and estrogen and progesterone receptor status, DDE was
found to be higher in women with breast cancer who were 50 years of age or older at diagnosis
and who had > 10 years between blood sample and diagnosis. The results, however, were not
statistically significant.
In a nested case-control study of 105 breast cancer cases and 208 matched controls
(subjects were participating in a prospective breast cancer study), risk ratios were found not to be
elevated for breast cancer in the highest quartile of lipid adjusted-serum DDE (Dorgan, 1999).
Breast cancer diagnoses were made up to 9.5 years after the collection of blood samples.
A case-control study examining postmenopausal women in western New York State (154
cases and 192 controls) found no association between breast cancer risk and current age- or
lipid-corrected serum DDE concentrations. Blood samples were collected within several months
of diagnosis between 1986 and 1991. Confounding factors considered included age, education,
familial breast cancer, parity, quetelet index (as a measure of obesity), age at first birth, duration
of lactation, years since last pregnancy, fruit and vegetable intake, and serum lipids (Moysich et
al., 1998).
Age-adjusted mean serum DDE levels (lipid-adjusted) in 475 case and 502 control
women in Connecticut were comparable (460 and 456 ppb, respectively), even when controlling
for menopausal status, parity, lactation, and by cases' estrogen receptor status. Patients (n=389)
with early stage disease had a slightly higher (not statistically significant) level of serum DDE
than the 20 patients with later stage disease (456 vs. 402 ppb, respectively; Zheng et al., 2000).
In Mexico City, DDT has been used in the recent past for malaria control. Between 1994
and 1996, hospital patients did not demonstrate a significant association between the level of
serum DDE (either wet weight or lipid weight basis using arithmetic or geometric mean) and
breast cancer risk. Cases consisted of 141 women with breast cancer. Age-matched control
DDE — January, 2008 7-7
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women were obtained from other areas in the hospital (excluding oncology and gynecology). No
association was found when menopausal status (using an odds ratio analysis adjusted for age),
quetelet index (as a measure of obesity), breast feeding with first birth, parity, family history of
breast cancer, and time elapsed since first birth were controlled (Lopez-Carrillo et al., 1997).
There was no association between current (1994) plasma DDE (or DDT) concentrations
(nonlipid-adjusted) and newly diagnosed breast cancer (n=21) in northern Vietnamese women.
The relative risk for breast cancer was not elevated significantly for subjects in the higher tertile
of plasma DDE, DDT, or total DDT (DDT plus molar-adjusted DDE) compared to the lowest
tertile. Women were similar with respect to age, age at menarche, age at first pregnancy, parity,
history of lactation, and maximum attained body weight (Schecter et al., 1997).
Mendonca et al (1999) found no association between breast cancer incidence and serum
DDE concentrations in women from Brazil. The study examined 177 cases of invasive breast
cancer and 350 controls from a large metropolitan region. Cases had a mean and median serum
DDE of 5.1 and 3.1 ppb, respectively, compared to 4.8 and 3.1 ppb, respectively, in controls.
The age-adjusted odds ratio for breast cancer in the upper quintile compared to the lowest
quintile was 0.90 (95% CI=0.47-1.73).
A study of DDE concentrations in breast adipose tissue from 73 breast cancer cases and
73 breast reduction surgery patients revealed that DDE levels were significantly elevated in
cases relative to controls; however, when adjusted for age the significance was lost (Bagga,
2000). Another case-control study of postmenopausal European women showed no significant
difference in current DDE concentrations in buttocks adipose for 265 women with breast cancer
(1.35 ppm) and 341 controls matched on age and study center (1.51 ppm; Van't Veer, 1997).
Laden et al. (2001a) performed a combined meta analysis from 5 studies (Helzlsouer et
al., 1999; Laden et al., 2001; Moysich et al., 1998; Wolff et al, 2000a; Zheng et al., 2000)
comparing women in the fifth quintile of lipid adjusted DDE values with the first quintile. The
multivariate pooled odds ratio for breast cancer was 0.99 (95% CI=0.77-1.27), indicating no
association between breast cancer risk and serum DDE concentrations. This analysis included
1400 case patients and 1642 controls.
Liljegren et al. (1998) found no association between malignant breast cancer and the
concentration of DDE in breast tissue fat. Forty-three cases of malignant breast cancer were
compared to 35 controls with benign breast disease. Fat samples were collected during surgery
after diagnosis from 1993-1995. Results did not change when the study groups were divided
into pre- and postmenopausal women or estrogen-receptor (ER-positive) subgroups. The
Liljegren et al. (1998) results were supported by Zheng et al. (1999), which also did not find a
significant association between breast adipose DDE concentrations and breast cancer. This
study included 304 cases of breast cancer and 186 benign breast disease controls in Connecticut.
The age-adjusted geometric mean of breast adipose DDE concentration was 737 and 784 in cases
and controls, respectively (DDT concentrations were 52 and 56 ppb, respectively).
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Hoppin et al. (2000) examined serum levels of DDE (as well as other organochlorine
compounds) in 108 subjects with pancreatic cancer and 82 controls (matched to cases by age and
sex) in a population-based study in the San Francisco Bay area. Cases had higher median lipid-
adjusted DDE concentrations (1.3 ppm) than the controls (1.0 ppm). The odds ratio for the
highest tertile compared to the lowest tertile was 2.1 (95% CI = 0.9-4.7), but it did not achieve
statistical significance. The odds ratio was reduced to 1.1 after controlling for PCB levels. As
part of this same study, Slebos et al. (2000) measured K-ras oncogene andp53 tumor suppressor
gene (two tumor-related molecular markers) and organochlorine levels in 61 subjects newly
diagnosed with pancreatic cancer. Patients with K-ras positive tumors tended to have lower
serum levels of DDE than those with K-ras negative tumors (significance not reported in
ATSDR, 2002), but there were no significant differences noted betweenp53 positive and
negative patients. Porta et al. (2000), however, determined that serum DDE (as well as DDT)
was significantly higher in 51 subjects with pancreatic cancer than in 26 controls (selection of
subjects not noted in ATSDR, 2002). The serum DDE levels (nonlipid adjusted) in the 17
patients with wild-type K-ras tumors did not differ from the control, however, the serum DDE
levels (nonlipid adjusted) in the 34 patients with mutated K-ras tumors were about double those
in controls. The study did not make comparisons using lipid-adjusted concentrations.
Hardell et al. (1996) did not find an association between DDE concentrations in
abdominal adipose tissue and non-Hodgkin's lymphoma. The mean DDE concentration in 28
newly diagnosed patients in Sweden was 1.4 ppm (lipid based) compared to 1.1 ppm in 17
surgical controls; however, nearly all of the 34 PCB congeners analyzed were significantly
higher in the cases relative to controls.
No association was found between lipid-corrected blood serum DDE concentration and
endometrial cancer in a multicenter case-control study of 180 women in the United States
(Sturgeon et al., 1998). Cases and controls both had median blood lipid-adjusted DDE
concentrations of approximately 1.4 ppm.
Cocco and Benichou (1998) used an ecological study design with multivariate statistical
techniques to examine the relationship between DDE concentration in subcutaneous fat, or levels
in tree bark (as a measure of environmental levels), and mortality from prostate and testicular
cancers in a number of States. The adipose samples for people in 22 states were collected from
the U.S. Department of Health, Education and Welfare's 1968 Human Monitoring Program; data
on DDE in tree bark were available for 18 states from 1992-1995. The age-adjusted mortality
rates from prostate and testicular cancers by state from 1971-1994 were obtained from the
National Center for Health Statistics. African Americans had a higher (74%) mean adipose DDE
concentration than whites, so separate analyses were performed for the two races. Numerous
demographic factors were considered possible confounders. The authors concluded that the
results did not support an association between prostate and testicular cancer and DDE exposure.
A separate study by Cocco et al. (2000) used the same methods discussed above, but
examined the association between DDE levels in adipose tissue (collected in 1968) and several
additional diseases. They examined DDE's association with age-adjusted mortality rates
between 1975 and 1994 for multiple myeloma, non-Hodgkin's lymphoma, and breast, uterus,
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liver, and pancreas cancer in 22 U.S. states. Separate analyses were conducted for gender and
race. No association was noted between DDE concentrations and pancreatic cancer or multiple
myeloma. There was an inverse correlation between adipose DDE and breast cancer mortality in
both white and African American women. An inverse correlation (statistically significant) also
was observed for uterine cancer in white women, but not in African American women. There
was no association observed for non-Hodgkin's lymphoma in either whites or African
Americans (men and women). Liver cancer mortality in white males and females was
significantly increased with adipose DDE concentrations (p<0.05), however, the same results
were not observed in African Americans. It should be noted that the number of subjects affects
the sensitivity of statistical analyses to find a significant effect.
7.2 Animal Studies
7.2.1 Acute Toxicity
LD50 studies of DDE are indicative of moderate toxicity. An LD50 of 880 mg/kg was
reported for male Sherman rats (Gaines, 1969). For o,p'-DDE 810 mg/kg was lethal in female
mice and 845.9 mg/kg in male mice (Tomatis et al., 1974).
As mentioned previously, DDE exposure is predominantly a result of its formation as a
metabolite of DDT. Accordingly, most studies of acute toxicity have examined the parent
compound, rather than the metabolite, and are not considered in this assessment.
7.2.2 Short-Term Studies
DDE mixed in corn oil and incorporated into the basal diet at concentrations of 316, 562,
1000, 1780, or 3160 ppm was administered to Osborne-Mendel rats (5/sex/group) for 6 weeks
followed by a 2-week observation period (NCI, 1978). These concentrations are equivalent to
doses of 0, 16, 28, 50, 89 or 158 mg/kg/day based on a food intake factor of 0.05 kg diet per kg
body weight (U.S. EPA, 1986d). This study was done as a range-finding exercise to help in
establishing the dose levels for the NCI cancer bioassay and did not examine a full complement
of standard toxicological endpoints. The only published version of the study results is in the
background section of the bioassay report, and include only information on the body weight and
mortality findings.
Body weight gain was reduced in male rats of all treatment groups. Male rats
administered a dose of 1000 ppm had body weights 11% lower than the control group and those
administered 1780 ppm were 22% lower than the control group. There were no exposure-related
changes in body weight for the females. One female rat administered 1000 ppm died, and all
female rats administered 1780 ppm or 3160 ppm died by the end of the 6-week treatment period.
The only deaths (number not specified) for the male rats occurred in the 3160-ppm group (NCI,
1978).
The same protocol was used with groups of B6C3F1 mice (5/sex/dose). DDE was mixed
in corn oil and incorporated into the basal diet at concentrations of 139, 193, 269, 373, or 519
DDE — January, 2008 7-10
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ppm (NCI, 1978). Using a food intake factor of 0.13 kg diet per kg body weight (U.S. EPA,
1986d), these concentrations are equivalent to doses of 0, 18, 25, 35, 47, and 68 mg/kg/day.
There were no changes in body weight associated with treatment. One control male and one
male in the 269-ppm dose group died. Four male mice and two female mice administered 373-
ppm dose group died (NCI, 1978).
Humoral (significantly increased serum albumin/globulin ratio, suppressed serum IgM,
IgG after ovalbumin immunization, and decreased antibody titer) and cell-mediated (increased
inhibition of leucocyte and macrophage migration, and decreased footpad thickness) immune
responses were adversely affected in male Wistar rats (8-12 per group). The effects occurred
after they were administered diets containing 200 ppm DDE, which was estimated to be
equivalent to 22.2 mg/kg/day, for 6 weeks. In addition, mean relative liver weight was increased
compared to the controls (Banerjee et al., 1996).
7.2.3 Subchronic Studies
No standard subchronic (13-week) oral studies for DDE were identified; however,
subchronic studies are available for DDT, the parent compound of DDE. A study of DDT
administered in the diet for 3 months suggests that the liver may be a target organ for DDE as
well as DDT. In this study a dose of 2.5 mg DDT/kg/day caused minor hepatocyte vacuolation
in male rats and a dose of 10 mg DDT/kg/day caused liver hypertrophy in females (Ortega,
1956). Chowdhury et al. (1990) administered 0.2 mg DDT/kg/day by gavage for 120 days to
rats. Atrophy of the adrenal gland in all zones except the zona glomerulosa was observed. In
male rats, the body weight gain was 30% lower in treated animals compared to the controls.
In another study, weanling rats (25/sex/group) were fed commercial DDT (8l%p,p'
isomer and 19% o,p' isomer) at levels of 0, 1, 5, 10 or 50 ppm for 15-27 weeks (Laug et al.,
1950). Increasing hepatocellular hypertrophy, especially in the centrilobular region, increased
cytoplasmic staining with an acid dye, and peripheral basophilic cytoplasmic granules were
observed at dose levels of 5 ppm and above. The response was dose-related becoming more
pronounced in the 10- and 50-ppm groups. No effects were reported for the 1-ppm concentration
(0.05 mg/kg/day), and this was identified as the no observed adverse effect level (NOAEL). The
authors stated that the effect seen at 5 ppm (0.25 mg/kg/day) seemed to represent the smallest
detectable morphologic effect, based on extensive observations of the rat liver as affected by a
variety of chemicals. The NOAEL from this study was used as the basis for the DDT RfD (U.S.
EPA, 1986e).
Affects on the immune response also have been observed after subchronic exposure to
DDT. Dietary administration of 1.9 or 19 mg DDT/kg/day for 31 days caused a decrease in the
severity of the anaphylactic response and the number of mast cells producing histamine after a
challenge with diphtheria toxoid compared to the control group (Gabliks et al., 1975). Adult
mice administered 0.06 or 0.63 mg DDT/kg/day for 16 weeks had a significant increase in the
DDE — January, 2008 7-11
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primary IgM response to sheep red blood cells (SRBC) and lymphoproliferative response to
LPS-coated (Escherichia coli lipopolysaccharide) SRBC; however, exposure for 20 or 24 weeks
caused a sharp reduction in both responses. A dose of 0.006 mg DDT/kg/day did not have an
effect (Rehana and Rao, 1992).
7.2.4 Neurotoxicity
No studies of DDE neurotoxicity are available; however, behavioral effects have been
observed in longer term bioassays. Minimal signs of neurotoxicity were reported in both the rats
and the mice during the National Toxicology Program (NTP) (1978) bioassay (full details of this
study are provided in Section 7.2.6 and 7.2.7). During the period of compound administration,
the male rats receiving 675- or 1350-ppm dietary concentrations (30 or 49 mg/kg/day) and
female rats receiving the 375- or 750-ppm concentrations (16 or 32 mg/kg/day) reportedly
displayed a hunched posture beginning after 8 weeks of treatment and continuing for the
remainder of the dosing period. The highest incidence occurred in the high-dose males. During
week 24, the dose levels were decreased and the incidence of hunched appearance decreased
dramatically in treated rats. Some of the control rats also exhibited a hunched posture during the
study but the incidence and severity were lower than that in the dosed rats.
Sixty to 85% of male B6C3F1 mice administered 150 or 300 ppm reportedly had a
hunched appearance after 22 weeks of treatment. The frequency and intensity of this effect
varied as dosing was altered within the same group during the study, strengthening its
association with dose. The doses within groups were not constant across the 78-week exposure
period as explained in Section 7.2.6. During the post-dosing observation period there was no
discernable difference between the exposed and control animals (NCI, 1978).
Rossi et al. (1983) conducted a chronic dietary exposure study (128 weeks) in groups of
40 to 47 male or female Syrian golden hamsters using concentrations of 0, 500 or 1000 ppm
DDE. No signs of neurotoxicity, such as tremors or convulsions were noted; the authors make
no reference to any possible changes in animal posture during the dosing period.
7.2.5 Developmental/Reproductive Toxicity
In the NCI (1978) chronic bioassay, DDE administered to Osborne-Mendel rats in the
diet at time-weighted-average concentrations of 0, 437, or 839 ppm for male rats or 0, 242, or
462 ppm for female rats for 78-weeks did not cause significant adverse effects on the ovaries,
uterus, mammary gland, testes, or prostate as revealed via routine histological examination of
these tissues. The same was true for B6C3F1 mice administered time-weighted-average
concentrations of 0, 148, or 261 ppm. There was, however, a dose-dependent, but not
statistically significant, increase in endometrial stromal polyps (0, 5, and 13%, respectively) in
female rats. Reproductive function was not evaluated in the NCI (1978) study.
DDE administered to rats for 5 weeks prior to mating and during gestation and lactation
at a dose of 10 mg/kg/day did not affect reproduction as measured by percent sperm positive,
DDE — January, 2008 7-12
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percent pregnant, gestation length, litter size, sex ratio of pups, and milk production and
composition (Kornbrust et al., 1986).
Reduced seminal vesicle and ventral prostate weights were observed when 200
mg/kg/day of DDE were administered for 4 or 5 consecutive days via gavage to male Long-
Evans rats (Kelce et al., 1995, 1997).
Exposure to 100 mg/kg/day during gestational days 14-18 in Sprague-Dawley rats
resulted in a significant decrease in ventral prostate weight in males at 15 months of age. Similar
exposure caused a decrease in weights of glans penis, ventral prostate, and epididymis in Long-
Evans males when they were 10 months old. Reduced anogenital distance and increased mean
number of retained nipples also were observed in the Long-Evans rats when they were newborn
(Grayetal., 1999).
Anogenital distance at birth was reduced in male Long-Evans pups exposed
transplacentally to 100 mg/kg/day during gestational days 14-18. Because of the long half-life
of DDE and its incorporation into milk, the prenatal exposure of the dams would have also led to
unquantified pup exposure during lactation. The animals also had retained thoracic nipples on
postnatal day 13. A significant delay in the onset of puberty (measured by the age of preputial
separation) was reported in male rats treated with 100 mg/kg/day from weaning (either day 21 or
25; ATSDR states that the specific day is unclear in text) until day 57 of age. Puberty was
delayed by 5 days, but was not accompanied by a change in serum testosterone levels (Kelce et
al., 1995, 1997).
A study by You et al. (1998) compared the effects of DDE on male sexual development
in offspring of Sprague-Dawley and Long-Evans rats. Gavage doses were administered to
pregnant dams at 10 or 100 mg/kg from gestation day 14 to 18. Reduced anogenital distance
developed in Long-Evans male rats, but the same effect was not noted in the Sprague-Dawley
rats. A dose of 10 mg/kg/day did not reduce the anogenital distance in males; anogenital
distance was not affected in females of either strain. Retention of thoracic nipples in Sprague-
Dawley pups was observed in the 10 mg/kg/day dose, but in Long-Evans pups this effect was
observed only with the 100 mg/kg/day dose. An apparent reduction of androgen receptor
expression (as measured by immunochemical staining) in male sex organs was noted in 100-
mg/kg/day Sprague-Dawley pups. Androgen receptor steady state messenger ribonucleic acid
(mRNA) was unaffected in Sprague-Dawley rats, but was increased 2-fold in Long-Evans rats
administered 100 mg/kg/day. The onset of puberty was unaffected in either strain with either
dose. The 10 mg/kg/day dose was an NOAEL in Long-Evans rats, but was the lowest observed
adverse effect level (LOAEL) in Sprague-Dawley rats.
Prenatal exposure to 10 and 100 mg/kg/day lead to nonsignificantly reduced prostate
ventral weights in males at 85 days of age, but did not affect the weights of the testes,
epididymis, or seminal vesicles. In addition, no change was noted in serum testosterone or LH
levels; however, DDE was associated with TRPM-2 (an androgen-repressed gene) expression
(Youetal., 1999a).
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Holtzman rats exposed during gestational days 14-18 with doses of 1 to 200 mg/kg/day
(exposure also occurred via mothers' milk from stored DDE) had a reduced anogenital distance
in males on postnatal day 1 and reduced relative ventral prostate weight on postnatal day 21 with
doses >50 mg/kg/day. Anogenital distance was still reduced on postnatal day 4 only in male
exposed to 200 mg/kg/day. Nipple retention on postnatal day 13 was significantly increased
with doses of 100 and 200 mg/kg/day. A delay in puberty of less than 2 days was noted in males
exposed to 200 mg/kg/day. On postnatal day 21, androgen receptor staining in the ventral
prostate was reduced in the 100 mg/kg/day dose, which was the only dose tested. Cauda
epididymal sperm number was reduced by 17% (relative to the control) on postnatal day 63 in
the 100 mg/kg/day dose group. Serum testosterone and 3a-diol androgens were not altered at
any time measured, nor were mRNA levels of androgen regulated genes from either the ventral
or dorsolateral prostate affected on postnatal day 21. The DDE body burden was not measured
at any point in the study. The LOAEL for this study was 50 mg/kg/day and the NOAEL was 10
mg/kg/day (Loeffler and Peterson, 1999).
7.2.6 Chronic Toxicity
In the NCI (1978) bioassay, DDE was administered to Osborne-Mendel rats in the diet at
time-weighted average concentrations of 0, 437, or 839 ppm in males and 0, 242, or 462 ppm in
females for 78 weeks. These concentrations are equivalent to approximately 0, 30, and 59
mg/kg/day respectively for males and 0, 16, and 32 mg/kg/day for females using the body weight
and food consumption values from U.S. EPA (1988ay. For the first 23 weeks of treatment (rats
were 7 weeks of age at the beginning of treatment), concentrations were 0, 675, or 1350 ppm in
males and 0, 375, or 750 ppm in females. Due to signs of toxicity the levels were reduced to 0,
338, or 675 ppm for males and 0, 187, or 375 ppm for females. In the low-dose group this
concentration was constant until week 78. In the high-dose group, this level was constant for 32
(females) or 36 (males) weeks; after this time, the high-dose group was cyclically administered
with 1 week free of treatment followed by 4 weeks of treatment (for a total of 15 weeks of
treatment and 4 weeks of no treatment for males and 18 weeks of treatment and 5 weeks of no
treatment for females). After 78 weeks of exposure (including the off weeks), rats were
observed for an additional 33-34 weeks. Body weight, food consumption, appearance, behavior,
clinical signs, and the incidence, size, and location of tumor masses were recorded. The skin,
subcutaneous tissue, lungs and bronchi, trachea, spleen, lymph nodes, thymus, heart, salivary
glands, liver, pancreas, esophagus, stomach, small and large intestines, pituitary, kidney, urinary
bladder, adrenal, thyroid, parathyroid, testis, prostate, brain, muscle, uterus, mammary gland,
ovaries and any tissue masses were routinely examined histologically.
The lowest dose administered caused a decrease in body weight gain in both males and
females. Mortality was 0, 16, and 28%, respectively, in females and 20, 32, and 48%,
1 U.S. EPA (1988a) does not provide food consumption information for Osborne Mendel rats. The values
used were those for mature Sprague Dawley rats because of similarities in average body weights and those of
Osborne Mendel rats. The body weights used for the dose calculation were estimated from the graphs of the average
body weights provided in the NCI (1978) report. Body weights over the second half of the study period provided the
basis for the body weight estimate.
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respectively, in males. Hunched or thin appearance was observed in dosed rats after 8 weeks of
treatment with the greatest incidence occurring in the high-dose males. During week 24, the
dose levels were decreased and the incidence of hunched appearance decreased dramatically, but
were still noted at a greater rate in treated rats than in the controls. Once DDE was completely
removed from the diet during week 78, control and treated rats had similar incidences of
hunched appearance. Fatty metamorphosis and centrilobular necrosis occurred in the liver of
both males and females with the incidence related to dose. There was no significant increase in
the incidence of any tumors (See Section 7.2.7). The LOAEL for this study was 16 mg/kg/day
for females and 30 mg/kg/day for males. The LOAEL was the lowest dose tested.
The NCI (1978) bioassay, also administered DDE to B6C3F1 mice in the diet at time-
weighted average concentrations of 0, 148, or 261 ppm in males and females for 78 weeks.
These doses were equivalent to approximately 0, 28, and 51 mg /kg/day in males, respectively,
and 0, 29, and 53 mg/kg/day in females, respectively. For the first 7 weeks of treatment (mice
were 6 weeks of age at the beginning of treatment) concentrations were 0, 125, or 250 ppm.
Concentrations were increased to 0, 150, or 300 ppm after 7 weeks since the previous doses were
well tolerated. These concentrations were administered for the remaining 71 weeks in the low-
dose group. These concentrations were constant in the high-dose group until week 33; after that
time the high-dose group was cyclically administered with 1 week free of treatment followed by
4 weeks of treatment for a total of 33 weeks on treatment and 9 weeks off treatment. After 78
weeks of treatment, the mice were observed for an additional 14 weeks. Body weight, food
consumption, appearance, behavior, signs of toxic effects, and the incidence, size, and location
of tumor masses were recorded. The skin, subcutaneous tissue, lungs and bronchi, trachea,
spleen, lymph nodes, thymus, heart, salivary glands, liver, pancreas, esophagus, stomach, gall
bladder, small and large intestines, pituitary, kidney, urinary bladder, adrenal, thyroid,
parathyroid, testis, prostate, brain, muscle, uterus, mammary gland, ovaries and any tissue
masses were routinely examined histologically.
There was a dose-dependent decrease in body weight in females as early as 10 weeks.
Although male body weights were similar to those for the controls throughout the study, final
body weights were decreased in a dose-dependent manner. Mortality was 5, 6, and 44%,
respectively, in females and 75, 30, and 38%, respectively, in males. Sixty to 85% of male
B6C3F1 mice administered 150 or 300 ppm reportedly had a hunched appearance after 22 weeks
of treatment. There was a dose-dependent increase in the incidence of chronic inflammation of
the kidneys in males. Low-dose females also had an increase incidence of chronic inflammation
of the kidneys. A significant increase in tumor incidence was noted (See Section 7.2.7). The
lowest dose tested (28 mg/kg/day for males and 29 mg/kg/day for females) was the LOAEL due
to effects on the kidneys and body weight gain (females). There was no NOAEL in the study.
Rossi et al. (1983) administered 500 or 1000 ppm DDE to groups of 40 to 47 male or
female Syrian golden hamsters in their diet beginning at 5 weeks and continuing until they were
128 weeks old. These concentrations are equivalent to 0, 56 or 122 mg/kg/day for males and 0,
51, or 114 mg/kg/day for females based on the food intake values for mature hamsters provided
in U.S. EPA (1988a) and average body weights from weeks 40 to 80 estimated from the graphic
presentations in the published paper. Survival did not appear to be affected by treatment
DDE — January, 2008 7-15
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although the numbers of males that died between weeks 50 to 100 in the high dose group was
greater than that for the controls. A dose-related reduction in body weight gain was reported for
the high dose males throughout the study and for the high dose females over weeks 50 to 80.
Body weight measurements were not made after week 80. Average body weights decreased
from weeks 50 to 80 for both males and females in the high dose group. Food consumption was
reported to be similar for all groups. In addition to an increase in liver tumors, hamsters
administered 1000 ppm developed focal or diffuse liver necrosis, fatty and cystic degeneration,
and hyperplastic foci of the liver. Liver necrosis, to a lesser extent, also was reported in 500-
ppm hamsters. An increase in adrenal tumors also was noted in male hamsters. The low-dose
group (56 mg/kg/day for males and 51 mg/kg/day for females) was an LOAEL for noncancer
effects based on the observation of liver necrosis.
Tomatis et al. (1974) administered 250 ppm DDE or a combination of 125 ppm DDE
with 125 ppm DDD via the diet to CF-1 (minimally inbred) mice from 6 to 7 weeks of age until
survivors were sacrificed at 130 weeks of age. These concentrations are equivalent to 0 or 33
mg/kg/day for DDE and 16 mg DDE/kg/day plus 16 mg DDD/kg/day 84 mg/kg/day for the
DDE/ODD mixture based on the food intake factors provided in U.S. EPA (1988a). Mortality
was greater in mice administered DDE alone compared to both the control and DDE plus DDD
group. All mice administered DDE died by 120 weeks. Survival in the DDE plus DDD group
was also lower than the control after 100 weeks. DDE and DDE plus DDD administration were
associated with lower body weights in males throughout the study. In females, only the DDE
plus DDD group had body weights lower than the control. Myocardial necrosis and diffuse
hemorrhage, leukocyte infiltration, and fibroblastic reaction were common in DDE treated mice
(22 of 60 males; 1 of 60 females). This also occurred in mice treated with DDE plus DDD, but
at a lower incidence (11 of 60 males) with no occurrence reported in the controls.
7.2.7 Carcinogenicity
In the NCI (1978) bioassay, DDE was administered to Osborne-Mendel rats in the diet at
time-weighted average concentrations of 0, 437, or 839 for males (0, 30, and 59 mg/kg/day) and
0, 242, or 462 ppm for females (0, 16, and 32 mg/kg/day). The methods of this study are
described in detail in the preceding Section 7.2.6. There was a slight, but not significant, dose-
related increase in thyroid tumors (follicular-cell adenomas and carcinomas in control, low-, and
high-dose groups were 11, 19, and 25%, respectively) in females. There was also a slight
increase in the males that was not related to dose (15, 24, and 21%, respectively). NCI,
however, did not consider there to be convincing evidence of carcinogenicity at these doses to
Osborne-Mendel rats.
The NCI (1978) bioassay, also administered DDE to B6C3F1 mice in the diet at time-
weighted average concentrations of 0, 148, or 261 ppm in males and females (equivalent to
approximately 0, 28, and 51 mg/kg/day in males and 0, 29, and 53 mg/kg/day in females). The
methods of this study are described in detail in the preceding Section 7.2.6. Hepatocellular
carcinomas occurred at a greater rate in treated mice (0, 17, and 36% in males, respectively; 0,
40, and 71% in females, respectively). The dose-related trend was statistically-significant in
DDE — January, 2008 7-16
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both males and females with high-dose males and both female treatment groups having a
significantly higher incidence compared to the control.
Rossi et al. (1983) administered 0, 500, or 1000 ppm DDE (0, 56 or 122 mg/kg/day for
males and 0, 51, or 114 mg/kg/day for females) to Syrian golden hamsters in their diet from 5
weeks of age until 128 weeks of age. Tumors were first observed in females after 28 weeks and
in males after 55 weeks. The first liver tumors did not appear until 76 weeks in the high dose
and 94 weeks for the low-dose females. The fist liver tumors did not appear until 105 weeks in
the low- and high-dose males. Table 7-1 provides a summary of the tumor data from this study.
Table 7-1 Tumors observed in the Rossi et al. (1983) study in Syrian Golden Hamsters
Dose
ppm
0
500
1000
Initial Animals
M
45
47
40
F
46
45
43
All Tumors*
M
15/31
11/30
20/39
F
13/42
13/39
16/39
Liver Tumors**
M
0/10
7/15
8/24
F
0/31
4/26
5/24
Adrenal Tumors*
M
8/31
5/30
17/39
F
2/42
7/39
8/39
* Based on number of animals when tumors first observed
** Based on the number of animals when liver tumors first observed
The total number of tumors and the number of hamsters with more than one tumor was
increased in 500-ppm females and 1000-ppm males and females. DDE treated hamsters were
found to have a greater incidence of liver tumors (0, 15.4, and 20.8%, respectively, in females; 0,
46.7, and 33.3%, respectively, in males). There was a dose-dependent increase in the average
number of liver nodules per animal and the average size of nodules accompanied by a decrease
in the time to first observation in females, but not in males. The liver tumors were described as
nodules on the surface of the liver that showed a loss of cellular architecture and compression of
the surrounding parenchyma. In addition to the tumors, eight of the high dose animals (3
females and 5 males) had hepatic hyperplastic foci. There was an increase in adrenal tumors as
well (5, 18, and 21%, respectively, in females; 26, 17, and 44%, respectively, in males), but the
results were not statistically significant.
Tomatis et al. (1974) administered 250 ppm DDE (33 mg/kg/day) or a combination of
125 ppm DDE with 125 ppm DDD (16 mg/kg/day of each compound) via the diet to CF-1
(minimally inbred) mice from 6 to 7 weeks of age until survivors were sacrificed at 130 weeks of
age. A statistically significant increase in incidence of hepatomas was observed in both males
and females in comparison with controls. In females, 98% of the 55 surviving exposed animals
developed hepatomas, compared to 1% of the surviving controls. In males, 74% of the animals
developed hepatomas as compared to 34% of the control animals. The appearance of the liver
tumors in CF-1 mice was similar to those associated with DDT exposure. Microscopically they
were either well-defined nodules, traebecular carcinomas or growths with mixtures of both
DDE —January, 2008
7-17
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characteristics. Some had pseudoglandular formations, particularly in the DDE-treated rats.
Four metastases to the lungs occurred in the DDE-treated female rats and two among the male
control rats.
7.3 Other Key Data
7.3.1 Mutagenicity and Genotoxicity
The direct studies of the mutagenicity and/or genotoxicity of DDE are limited. DDE was
mutagenic in mouse lymphoma (L5178Y) cells and Chinese hamster (V79) cells, but was
negative in Salmonella (ICPEMC, 1984). Testicular DNA synthesis was not inhibited in male
mice administered a single oral dose of 50 mg DDE/kg (Seller, 1977).
DDE was reported to induce chromosomal damage in vitro in the B14F28 Chinese
hamster cell line (Mahr and Miltenburger, 1976). Chinese hamster V79 cells exposed to DDE
with activation had a significant increase in chromosomal aberrations. DDT did not cause the
same effect (Kelly-Garvert and Legator, 1973). Positive results for chromosomal aberrations
were also reported for kangaroo rat (Potorus tridactylis) cells by Palmer et al. (1972).
NTP (2005) reported mixed results in genetic toxicity testing assays (Table 7-2).
Table 7-2 Summary of NTP (2005) Genetic Toxicology Results
in vitro Cytogenetics
Drosophila
Mouse Lymphoma
Salmonella
Negative (chromosome aberrations)
Weakly Positive (sister chromatid exchanges)
Positive (sex-linked recessive lethal)
Negative (reciprocal translocation)
Positive
Negative (2 tests)
7.3.2 Immunotoxicity
Humoral and cell-mediated immune responses were adversely affected in male Wistar
rats (8-12 per group) administered diets containing 200 ppm DDE, which was estimated to be
equivalent to 22.2 mg/kg/day, for 6 weeks. Humoral responses included significantly increased
serum albumin/globulin ratio, suppressed serum IgM, IgG after ovalbumin immunization, and
decreased antibody titer. Cell-mediated responses included increased inhibition of leucocyte and
macrophage migration, and decreased footpad thickness (Banerjee et al., 1996).
NCI (1978; see Section 7.2.6 for study description) showed no treatment-related adverse
effects on the thymus, spleen or lymph nodes in either rats or mice. Immunocompetence was not
evaluated in this study.
DDE — January, 2008 7-18
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7.3.3 Hormonal Disruption
DDE has been shown to have weak estrogenic activity when tested at doses up to 0.5 mM
in yeast expressing the human estrogen receptor, MCF-7 and HeLa cells. o,p'-DDE was slightly
more potent thanp,p'-DDE, but both were approximately 106 times less potent than l?p-estradiol
on a molar basis (Balaguer et al., 1999; Sohoni and Sumpter, 1998; Tully et al., 2000). The in
vitro E-screen test indicated that/\p'-DDE was a partial estrogenic agonist, but that 107 times
more DDE was needed to produce maximal results as compared to l?p-estradiol (Soto et al.,
1998). Using yeast, Gaido et al. (1997) found o,//-DDE produced little response in expression of
the estrogen receptor, a reporter gene regulated by two estrogen receptor response elements, or a
yeast gene that is supposed to enhance steroid receptor mediated transcription.
It has been shown that o,p'-DDE competes with l?p-estradiol for binding to the estrogen
receptor in rabbit uterine extracts (concentration that inhibited specific binding by 50% =
[IC]50=40 |IM for DDE; IC50=2.7 nM for l?p-estradiol; Danzo, 1997) and immature rats (IC50=5
|IM for DDE; IC50=0.8 nM for 17p-estradiol; Kelce et al., 1995). The/y/-DDE isomer was
found to be a relatively ineffective competitor, which is consistent with the observation of earlier
in vitro studies (Hitman and Cecil, 1970; Gellert et al., 1972; Nelson, 1974).
The interaction of DDE with the androgen receptor has also been studied. In a
competitive androgen receptor binding assay in rat ventral prostate cytosol using a radiolabeled
synthetic androgen (R1881), DDE had an inhibition constant of 3.5 |lM, which was similar to
that of diethylstilbesterol (DES) and about 30 times weaker than 17p-estradiol. DDE bound to
the androgen receptor 200 times more efficiently than the estrogen receptor (in uterine cytosolic
extracts from immature rats) (Kelce et al., 1995).
DDE was 20-fold less effective than DES or 17p-estradiol in inhibiting the conversion of
testosterone by 5a-reductase to 5a-dihydrotestosterone and 5a-androstan-3a-17p-diol in
microsomes isolated from the adult rat caput and corpus epididymis. The study also confirmed
that DDE (IC50=6.8 |lM) could compete with dihydrotestosterone (IC50=1.1 nM) for binding to
androgen receptors in rat prostate cytosol. The study, however, also found o,//-DDE to only be
slightly less effective than/\p'-DDE in competing with dihydrotestosterone for androgen
receptor binding (Danzo, 1997).
Castrated adult (120 days old) male rats implanted with testosterone-containing Silastic
capsules to maintain a constant serum testosterone level were administered 200 mg/kg/day DDE
for 4 days. A significant reduction in androgen-dependent seminal vesicle and ventral prostate
weight relative to the control were reported. Prostates from treated rats had a 13-fold increase in
androgen-repressed testosterone-repressed prostatic message 2 (TRPM-2) messenger RNA levels
and a 35% decline in androgen-induced prostate binding subunit 3 (C3) mRNA levels relative to
the control (Kelce et al., 1995). Similar results were reported by Kelce et al. (1997) after 5 days
of administering the same dose of DDE; however, they reported that testosterone metabolism
was not affected.
DDE — January, 2008 1-19
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Immunohistochemical staining of androgen receptor in epididymal nuclei of adult rats
administered 200 mg/kg/day for 5 days was significantly reduced compared to the control.
TRPM-2 was significantly increased and testosterone-induced C3 was significantly decreased.
In addition, the weights of the seminal vesicle and ventral prostate weight were significantly
reduced (Kelce et al., 1997).
Kelce et al. (1995) transiently transvected monkey kidney cells with an androgen
receptor expression vector and a reporter gene containing the MMTV promoter, which contains
binding sites for the androgen receptor. Both 0.2 |lM/\p'-DDE and 1 |lM hydroxyflutamide
inhibited 5a-dihydrotestosterone (0.1 nM) induced transcription by about 50%. In contrast, in
lysates of rat ventral prostate, the androgen antagonist hydroxyflutamide was 10 times more
effective than p,p'-DDE in inhibiting binding to the androgen receptor. The authors suggested
that their findings raised the possibility that the androgen receptor, rather than the estrogen
receptor, is the site of hormonal blockade by persistent environmental pollutants such asp,p'-
DDE.
DDE has been found to inhibit in vitro androgen receptor regulated gene expression in
HEPG2 human hepatoma cells transiently transected with the human androgen receptor and a
reporter gene linked to an androgen responsive promotor. These cells were exposed for 24 hours
to dihydrotestosterone (up to 0.1 |lM) with various doses of DDE (doses not specified in
ATSDR, 2002). Graphical presentation of the data indicates that both DDE isomers were
equipotent in inhibiting dihydrotestosterone induced gene transcription. The authors state that
DDE was the most potent inhibitor (DDT and DDD also were tested) with an IC50 of 1.86 |lM,
but statistical significance is not mentioned. The study did not sample the cells or media to
determine if the metabolism of dihydrotestosterone was affected (Maness et al., 1998).
DDE was shown to inhibit dihydrotestosterone responsive gene expression in yeast
expressing the human androgen receptor plus a secreted reported gene controlled by androgen
response elements. A 4-day incubation with 1.25 nM dihydrotestosterone (a submaximally
inducing dose) appeared to have an IC50 value of approximately 10 |iM; however, after 5 or 6
days of incubation dihydrotestosterone appeared to partially overcome the inhibition. The effect
of DDE on dihydrotestosterone metabolism was not examined (Sohoni and Sumpter, 1998).
Based on these results, DDE has been demonstrated as an androgen receptor antagonist
by its ability to inhibit the androgen receptor from either appropriately inducing or repressing
transcription from androgen responsive genes (Kelce et al., 1995, 1997; Maness et al., 1998;
Sohoni and Sumpter, 1998). The androgen antagonist mechanism demonstrated in these studies
would explain a number of reproductive and developmental effects seen in male rats exposed to
DDE at various ages.
7.3.4 Physiological or Mechanistic Studies
DDE, like DDT, seems to cause to liver toxicity. Fatty metamorphosis and centrilobular
necrosis were seen in both male and female rats in the NCI (1978) study. Male and female mice
had a statistically significant, dose-related increase in the incidence of hepatic tumors in this
DDE — January, 2008 7-20
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same study. In concert with these observations, there were several studies of liver enzymes,
including cytochrome P450s that illustrate biochemical responses to DDE exposure.
Nims et al. (1998) observed increases in CYP2B, CYP3A1 and CYP3A2, which were
accompanied by a significant increase in relative liver weight in rats administered DDE in the
diet (0.15 to 36 mg/kg/day) for 14 days. The NOEL for CYP2B induction was 0.17 mg/kg/day,
while the liver weight was significantly increased with a dose of >4.1 mg/kg/day. Rats exposed
to two gavage doses of 350 mg DDE/kg exhibited an increase in ornithine decarboxylase and
cytochrome P-450 levels (Kitchin and Brown, 1988). CF1 mice administered 42.9 mg/kg/day
via gavage for 7 days had increased liver weight (29% increase in absolute liver weight),
microsomal P450, cytochrome-C reductase, and serum total protein (Pasha, 1981). Adult rats
administered 100 mg/kg/day for 7 days had elevated hepatic aromatase (an enzyme involved in
steroid metabolism; You et al., 2001). The most sensitive of these responses is the induction of
CYP2B at a dose of 0.17 mg/kg/day. The CYP2B family is involved in the metabolism of
xenobiotic compounds.
You et al. (1999b) examined the potential of in utero DDE exposure to affect the
developmental expression of hepatic CYPs enzymes responsible for testosterone hydroxylation.
Pregnant Sprague-Dawley rats were treated with DDE (0, 10, or 100 mg/kg/day) on gestational
days 14-18. Adult male rats also were treated for comparison. The results indicate that
responses of CYP2C11 and CYP2A1 are regulated differently in relationship to developmental
stages. DDE induced 2A1 in males on postnatal day 10, but not on postnatal day 21.
Pronounced induction of 2B1 was observed in both males and females on postnatal days 10 and
21. 3A1 was induced to a lesser extent, and 2C11 was not induced. DDE induced 2B1, 3A1,
and 2C11 in the adult males, but 2A1 was not induced. You et al. (2001) found that aromatase,
which catalyzes the conversion of C19 steroids to estrogen, was increased in hepatic microsomes
of treated adult male rats, as were the levels of aromatase protein in the liver. Adult rats
administered 100 mg/kg/day for 7 days had elevated hepatic aromatase (an enzyme involved in
steroid metabolism), which is a potential mechanism by which DDE could affect reproduction
and/or sexual maturation (You et al., 2001).
In cultures of ER-positive MCF-7 human breast cancer cells exposed to o,p'-DDE (dose
not specified in ATSDR, 2002), the ratio of estradiol metabolites 16a-OHEl/2-OHEl was
increased over controls. It also was found that E22-hydroxylation was reduced compared to the
control group. It was noted, however, that the ratio did not consistently predict known mammary
carcinogens in the same assay. o,p'-DDE (dose not specified in ATSDR, 2002) did not induce
significant changes in either C2 or 16a-hydroylation of estradiol in ER-negative MCF-10 or
MDA-MB-231 cell lines, suggesting an estrogenic mechanism pathway (Bradlow et al., 1995,
1997; McDougal and Safe, 1998).
As discussed above (Section 7.3.3), the mechanism DDE's endocrine action appears to be
via antagonism at the androgen receptor. This mechanism could explain a number of
reproductive and developmental effects seen in male rats exposed to DDE at various ages.
DDE — January, 2008 7-21
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7.3.5 Structure-Activity Relationship
No studies were identified that examined the structure-activity relationship of DDE and
other aromatic organochlorine compounds. However, DDT, DDE and ODD are related
compounds that differ only in their degree of chlorination or saturation. DDT differs from DDD
in that it has three chlorine atoms bound to the terminal carbon on the aliphatic potion of the
molecule in place of two, and DDE differs from DDD in that the aliphatic side chain is
unsaturated (includes a double bond). DDE and DDD are degradation products and metabolites
of DDT. According, it is likely that the three compounds will have a number of properties in
common.
7.4 Hazard Characterization
7.4.1 Synthesis and Evaluation of Major Noncancer Effects
DDE has been demonstrated to reduce body weight gain in Osborne-Mendel rats,
B6C3F1 mice, CF-1 mice and Syrian hamsters. While neurological effects were noted in both
rats and mice, they were not noted in the hamster. In rats and hamsters, the liver was a target
organ. Rats developed fatty metamorphosis and centrilobular necrosis in the livers of both males
and females with the incidence related to dose. Hamsters developed focal or diffuse liver
necrosis, fatty and cystic degeneration, and foci of hyperplastic cells of the liver. In mice, the
kidney was a target organ. There was a dose-dependent increase in the incidence of chronic
inflammation of the kidneys in males; low-dose females also had an increase incidence of
chronic inflammation of the kidneys (there was greater than 7-fold higher mortality in the high-
dose females, which may be related to the lack of findings in this group).
DDE has been found to be an antiandrogenic compound, which may explain a number of
reproductive and developmental effects seen in male rats exposed to DDE at various ages.
Humoral and cell-mediated immune responses were adversely affected in male Wistar rats (22.2
mg/kg/day, for 6 weeks), and subjects with higher blood DDE levels were demonstrated to have
lower mitogen (concanavalin A)-induced lymphoproliferative activity and slightly increased
total lymphocytes immunoglobulin A levels have been observed as immune-related effects.
The potential influence of DDE body burden on health effects is not well understood, but
is an important consideration given the potential for accumulation in lipid tissues. Similarly, the
health effects associated with the combined exposure to DDT, DDE, and other related
metabolites have not been addressed.
7.4.2 Synthesis and Evaluation of Carcinogenic Effects
In animals treated with DDE via the diet, increases in the incidence of liver tumors,
including carcinomas, were observed in two strains of mice and in hamsters; there were also
nonsignificant increases in the incidence of thyroid tumors in female rats and adrenal tumors in
hamsters.
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Human epidemiology studies have not established any conclusive link between any
specific type of cancer and DDE exposure. Although there have been several studies that
observed an association between DDE levels in serum or adipose tissues and breast cancer, there
are numerous other studies that did not find such an association.
7.4.3 Mode of Action and Implications in Cancer Assessment
DDE has been demonstrated to cause chromosomal damage or aberrations in several cell
types but was only mutagenic in the mouse lymphoma and Chinese hamster cell assays. There
also is some evidence that DDE alters the immune system, which is a feature that can facilitate
the tumorigenic process. However, a mode of action and sequence of key events in DDE
carcinogenesis has not been established.
7.4.4 Weight of Evidence Evaluation for Carcinogenicity
DDE is likely to be carcinogenic to humans. This conclusion is based on increases in the
incidence of liver tumors, including carcinomas, in two strains of mice (B6C3F1 and CF-1) and
in hamsters (NCI, 1978; Rossi et al., 1983; Tomatis et al., 1974). The incidence of thyroid
tumors was increased in female rats after dietary exposure to DDE (NCI, 1978); however, the
NTP did not consider the increase to be significant and concluded that the evidence did not
support carcinogenicity in male and female Osborne-Mendel rats. The evidence of
carcinogenicity from human epidemiology studies is inconclusive.
7.4.5 Potentially Sensitive Populations
Potential Gender Sensitivity
In the cancer bioassay (NCI, 1978) females had a stronger tumor (liver carcinomas in
mice, thyroid follicular-cell adenomas and carcinomas in rats) response to dose than did the
males. In the short-term study by NCI (1978), males were more susceptible to body weight
effects, and females to mortality, after exposure to DDE.
DDE's potency as an antiandrogenic agent is stronger than its estrogenic properties. This
androgen antagonism or antiandrogenic activity could explain a number of reproductive and
developmental effects seen in male rats of various ages. Observed effects in the male animals
include reduced anogenital distance and retention of thoracic nipples in pups exposed during
gestation and lactation (Gray et al., 1999; Kelce et al., 1995; Loeffler and Peterson, 1999; You et
al., 1998); delayed puberty in rats exposed either during juvenile development (Kelce et al.,
1995) or at very high doses during gestation and lactation (Loeffler and Peterson, 1999); and
reduced accessory sex organ weights in exposed adult males (Gray et al., 1999; Kelce et al.,
1995, 1997; You et al., 1999a).
Neonates, Infants, and Fetuses
There is some evidence in humans that DDE may have adverse effects on reproductive
outcome. Prenatal exposure may affect parturition timing, as observed in two epidemiologic
studies. The levels of DDE found in the blood of women who had early deliveries were higher
DDE — January, 2008 7-23
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than that for women who delivered at term (O'Leary et al., 1970a). Quadratic spline models
were used to examine 2830 birth records relative to maternal serum levels of DDE and found
that the odds of pre-term birth began to increase at a DDE concentration of 10 ppb (Longnecker
etal.,2001).
Due to its high Kow, DDE selectively partitions into human breast milk. Infants exposed
to DDE in maternal milk at levels >4 ppm had poorer reflect reactions than children exposed to
lower levels (Rogan et al., 1986, 1987). The same authors found an inverse relationship between
the levels of DDE in maternal milk and lactation duration in women in the United States and
Mexico. Gladen and Rogan (1995) examined the relationship between DDE and lactation
duration in Mexico and found that the inverse relationship between DDE and duration of
lactation only applied to women with prior breast-feeding experience. ZDDT concentrations
(which includes DDE) in human breast milk have been declining steadily across the U.S.,
Canada, and Western Europe (approximately one half in 4.2-5.6 years for ZDDT; ATSDR,
2002).
Some of the epidemiological studies discussed in Section 7.1.2 have suggested that DDE
may adversely affect growth in children. Serum samples collected from 2380 pregnant women
with a median DDE concentration of 25 ppb (range 3-178 ppm), showed that the adjusted odds
for small-for-gestational-age births was increased at a blood DDE concentration of 10 ppb
(Longnecker et al., 2001). Karmaus et al. (2002) found that girls between 1.3 months and 8
years of age exposed to 0.44-40.4 ppb of DDE were almost 2 cm shorter than those exposed to
lower concentrations; the decrement in height was not observed in these same girls at age 10.
Prenatal exposure to DDE has also been associated with increased height and weight in boys at
puberty (Gladen et al., 2000).
As discussed in Section 7.2.5, animal studies also reveal susceptibility to DDE during
development, particularly in males. In rats, exposure to 100 mg/kg/day during gestation resulted
in a significant decrease in ventral prostate weight in males at 15 months of age and a decrease in
weights of glans penis, ventral prostate, and epididymis at 10 months old; reduced anogenital
distance and increased mean number of retained nipples also were observed in the newborns
(Gray et al., 1999). Anogenital distance at birth was reduced in male rat pups exposed
transplacentally to 100 mg/kg/day during gestational (and also unquantified pup exposure during
lactation); the animals also had retained thoracic nipples on postnatal day 13 and a significant
delay in the onset of puberty was reported (Kelce et al., 1995, 1997).
You (1998) administered doses of 10 or 100 mg/kg to Long-Evans and Sprague-Dawley
rats from gestation day 14 to 18. Reduced anogenital distance developed in the Long-Evans
male rat. Retention of thoracic nipples in Sprague-Dawley pups was observed in the 10
mg/kg/day dose, but in Long-Evans pups this effect was observed only with the 100 mg/kg/day
dose. An apparent reduction of androgen receptor expression (as measured by immunochemical
staining) in male sex organs was noted in 100 mg/kg/day Sprague-Dawley pups; androgen
receptor steady state messenger ribonucleic acid (mRNA) was increased 2-fold in the Long-
Evans. Prenatal exposure to 10 or 100 mg/kg/day lead to nonsignificantly reduced prostate
DDE — January, 2008 7-24
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ventral weights in males at 85 days of age and was associated with TRPM-2 (an androgen-
repressed gene) expression (You et al., 1999a).
Holtzman rats exposed during gestational days 14-18 with doses of 1 to 200 mg/kg/day
(exposure also occurred via mothers' milk from stored DDE) had a reduced anogenital distance
in males on postnatal day 1 and 4, reduced relative ventral prostate weight on postnatal day 21,
significantly increased nipple retention on postnatal day 13, and delayed puberty (of less than 2
days in males). On postnatal day 21, androgen receptor staining in the ventral prostate was
reduced, and Cauda epididymal sperm number was reduced on postnatal day 63 (Loeffler and
Peterson, 1999).
Early-life exposure to DDT, the parent compound of DDE, has been associated with an
increased tumor incidence (U.S. EPA, 2005b). Vesselinovitch et al. (1979) assessed early-life
exposures to DDT and resulting liver tumors in B6C3FJ mice. The analysis performed by the
U.S. EPA in the Supplemental Guidance for Assessing Susceptibility from Early Life Exposure to
Carcinogens found that the median ratio of juvenile to adult potency for DDT is 2.5 (male mice;
U.S. EPA, 2005b). Similar data are not available for DDE in order to assess the potential for
early-life exposure effects. The Supplemental Guidance did not establish default adjustments (to
be used when chemical-specific data are not available) for chemicals with nonmutagenic modes
of action, such as DDT.
DDE — January, 2008 7-25
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8.0 DOSE-RESPONSE ASSESSMENT
8.1 Dose-Response for Noncancer Effects
A limited number of conventional dose-response studies are available to assess the dose-
response of DDE. Limited data on DDE, mostly from the NCI (1978) bioassay, suggest that the
liver is a target organ in mammalian species. Hormonal effects on reproduction and sexual
development are also important aspects of DDE toxicity. A reference dose (RfD) has not been
developed for DDE; however, there is an RfD for the parent pesticide DDT of 0.0005 mg/kg/day
(based on an NOAEL of 0.05 mg/kg/day from a dietary subchronic study; U.S. EPA, 1986e). In
this study, liver lesions were identified at an LOAEL of 0.25 mg/kg/day. This RfD has relevance
to DDE because: 1) DDE is the principle persistent metabolite of DDT; and 2) both compounds
have a number of similar effects. An RfD is estimate (with uncertainty spanning perhaps an
order of magnitude) of a daily oral exposure to the human population (including sensitive
subgroups) that is likely to be without an appreciable risk of deleterious effects during a lifetime.
ATSDR has established acute and intermediate-term Minimum Risk Levels (MRLs) for
DDT's noncancer effects (ATSDR, 2002). The DDT oral acute MRL is 0.0005 mg/kg/day,
based on developmental effects and applying and uncertainty factor of 1000. The in oral
intermediate-term MRL for DDT is also 0.0005 mg/kg/day, but is based on hepatic effects and
the application of a 100-fold uncertainty factor.
The World Health Organization (2004) recently established a Provisional Tolerable Daily
Intake (TDI) value for DDT and its derivatives of 0.01 mg/kg/day based on an NOAEL of 1
mg/kg/day for developmental effects in rats and applying a 100-fold uncertainty factor.
8.2 Dose-Response for Cancer Effects
8.2.1 Choice of Study
There are three studies that demonstrate the carcinogenic potential of DDE in animals.
Table 8-1 provides a summary of the dose and response information from these three studies.
The NCI (1978) study was chosen as the principal study because it provided the best dose-
response data and because the mice appear to be more sensitive to the carcinogenic effects of
DDE than the hamster.
The NCI (1978) bioassay, administered DDE to B6C3F1 mice in the diet at time-
weighted average concentrations of 0, 148, or 261 ppm in males and females (equivalent to
approximately 0, 28, and 51 mg/kg/day in males and 0, 29, and 53 mg/kg/day in females). The
methods of this study are described in detail in the preceding Section 7.2.6. Hepatocellular
carcinomas occurred at a greater rate in treated mice (0, 17, and 36% in males, respectively; 0,
40, and 71% in females, respectively). The dose-related trend was statistically-significant in
both males and females, with high-dose males and both female treatment groups having a
significantly higher incidence compared to the control.
DDE — January, 2008 8-1
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Table 8-1 Dose Response Data for the Carcinogenicity of DDE
Administered
Concentration
mg/kg/day1
Tumor Incidence
female
male
Reference
Mouse/B6C3Fl; hepatocellular carcinomas
0
28M/29F
51M/53F
0/19
19/47
34/48
0/19
7/41
17/47
NCI, 1978
Mouse/CF-1; hepatomas
0
33
1/90
54/55
33/98
39/53
Tomatis et
al., 1974
Hamsters/Syrian Golden; neoplastic nodules (hepatomas)
0
54
118
0/31
4/26
5/24
0/42
7/15
8/24
Rossi et
al., 1983
1. Doses derived from experimental body weights and food intake values from U.S. EPA (1988a)
Tomatis et al. (1974) administered 250 ppm DDE (33 mg/kg/day) via the diet to CF-1
mice from 6 to 7 weeks of age until survivors were sacrificed at 130 weeks of age. A
statistically significant increase in incidence of hepatomas was observed in both males and
females in comparison with controls. In females, 98% of the 55 surviving exposed animals
developed hepatomas, compared to 1% of the surviving controls, however, this study used only
one dose and is not appropriate for dose-response modeling.
Rossi et al. (1983) administered 0, 500, or 1000 ppm DDE (0, 54, or 118 mg/kg/day
average doses) to Syrian hamsters in their diet from 5 weeks of age until 128 weeks of age.
Although the number of tumor-bearing hamsters was the same regardless of treatment, the total
number of tumors and the number of hamsters with more than one tumor was increased in 500-
ppm females and 1000-ppm males and females. DDE treated hamsters were found to have a
greater incidence of liver tumors (0, 15.4, and 20.8%, respectively, in females; 0, 46.7, and
33.3%, respectively, in males). There was a dose-dependent increase in the average number of
liver nodules per animal and the average size of nodules accompanied by a decrease in the time
to first observation in females, but not in males. There was an increase in adrenal tumors as well
(5, 18, and 21%, respectively, in females; 26, 17, and 44%, respectively, in males), but the
results were not statistically significant.
For the current assessment, the tumor data from the mice in the NCI (1978) study were
used for quantification. Both the NCI (1978) and the Rossi et al (1983) study in hamsters
included a control and two dose groups. However, the mice appeared to be more sensitive to the
tumorigenic effects of DDE than hamsters. The Tomatis et al. (1974) study could not be used
because it employs a single dose and, therefore, cannot be used for dose-response modeling
following to the U.S. EPA (2005a) guidelines for carcinogen risk assessment. The hepatomas
observed in the hamsters are important to the risk assessment because they demonstrate that the
carcinogenic effects of DDE are not limited to mice.
DDE —January, 2008
8-2
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8.2.2 Dose-Response Characterization
Groups of male and female B6C3F1 mice (50/sex/treated mice and 20/sex/control group)
were administered commercially available DDE in corn oil incorporated into the diet at time-
weighted concentrations of 0, 148 and 261 ppm. Normalized-time-weighted-average (TWA)
daily doses were approximately 0, 28 and 51 mg/kg/day, respectively, in males and 0, 29, and 53
mg/kg/day, respectively, in females. Effects in treated animals were compared to the untreated
control animals. A highly significant dose-related trend in the incidence of hepatocellular
carcinomas was observed in males and females that received the chemical when compared to the
respective controls (Table 8-2). These tumors also appeared earlier in mice administered the
higher dose. Slightly decreased body weight gain and increased mortality also were observed in
exposed female mice, but were not observed in the male mice. There were no apparent increases
in the incidences of non-neoplastic lesions (NCI, 1978).
Table 8-2 Summary of Liver Tumor Incidence, 78-Week Study in Mice
Male
Female
Dietary Concentration
(ppm)
0
148
261
0
148
261
Normalized-time-weighted-averagedose
(mg/kg/day)
0
22
51
0
29
53
Liver Tumor Incidence
0/19
7/41
17/47*
0/19
19/47*
34/48*
Source: NCI (1978)
* Significantly increased compared to control, p<0.05
8.2.3 Extrapolation Models and Rationale
The data from the NCI (1978) study in mice were analyzed using the EPA Benchmark
Dose Software (BMDS) Version 1.3.2, and results are summarized in Appendix B. For female
mice, the multistage (2) model, extra risk, was used, yielding an ED10 of 8.09 mg/kg-day and a
LED10 of 4.19 mg/kg-day; the LED10 was selected as the point of departure for the cancer
analysis and a linear extrapolation was applied. A linear low-dose extrapolation is appropriate
for the assessment of a carcinogen lacking a demonstrated mode-of-action. Figure 8-1 shows the
fit of the multistage (2) model to the data for the female mice.
DDE —January, 2008
8-3
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Figure 8-1 Multistage (2) Model with 0.95 Confidence Interval for the Female Mice
Data from the NCI (1978) Tumor Bioassay
Multistage Model with 0.95 Confidence Level
T3
•§
t
<
f—
O
'"C
CD
^
Ll_
0.8
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0
ivuiuoieiLjc; ;
BMD Lower Bound :
^-^ i
„/ :
/^ j
.s^ ~_
A- // \
// // ;
/ ^^ !
^^ ^
y _^^
/^^ !
... BMDL , , , BMD , , , , i
10
20
30
40
50
dose
13:5703/232005
8.2.4 Cancer Potency and Unit Risk
Oral
The NCI (1978) female mouse hepatocellular carcinomas data were BMD modeled,
yielding an ED10 of 8.09 mg/kg-day and LED10 of 4.19 mg/kg-day, which were used as the points
of departure in the derivation of central tendency and lower bound slope factors for DDE. The
rodent ED10 and LED10 values were converted to human equivalent doses (HED) by applying a
3/4-power body weight adjustment, as recommended by the U.S. EPA Guidelines for Carcinogen
Risk Assessment Guidelines (U.S. EPA, 2005a). This resulted in HED ED10 and HED LED10
values of 1.16 and 0.60 mg/kg/day, respectively. The central tendency and lower bound slope
factor estimates for carcinogenicity are calculated from these values as follows:
Central Tendency Estimate:
Slope Factor (SF) = Response = 0.1
= 8.6x 10'2 (mg/kg-day)-
ED
10
Lower Bound Estimate:
1.16 mg/kg/day
0.1
Slope Factor = Response =
LED10 0.60 mg/kg/day
= 1.7x ID'1 (mg/kg-day)-1
DDE —January, 2008
8-4
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The Health Reference Level (HRL) serves as the benchmark for examining the
occurrence data for DDE in the Regulatory Determination process. It is the concentration in
drinking water equivalent to a one-in-a million risk (10"6) of cancer above back ground. For
DDE, the 10"6 risk is calculated as follows:
10"6risk= risk x body weight
SF x drinking water intake
0.000001x70 kg
= 2.1x 10-4mg/L
0.17 (mg/kg/day)'1 x 2L/day
The FIRL is rounded to one significant figure and becomes 0.2 |ig/L
Prior Cancer Slope Factor
EPA evaluated the carcinogenicity for DDE under the Guidelines for Cancer Risk
Assessment (U.S. EPA, 1986c) using the linear multistage model. The oral slope factor using
this approach is 3.4 x 10"1 per (mg/kg-day) using administered doses that were adjusted for
frequency and length of exposure, and human equivalent dose (U.S. EPA, 1988b; see Table 8-3).
The FIRL calculated from the IRIS slope factor is 0.1 |ig/L.
Table 8-3 Factors Used to Derive the Previous Oral Slope Factor for DDE
Administered Dose
(ppm)
0
148
261
Human Equivalent Dose
(mg/kg-day)
0
0.9
1.58
Tumor Incidence
(Female mice)
0/19
19/47
34/48
Source: U.S. EPA (1988b)
The U.S. EPA (1988b) assessment on IRIS differs from the revised assessment presented
in this document in a number of respects. The differences in the approach are summarized as
follows:
• The 1988 HED was determined using a body weight raised to the 2/3-power as
recommended by the EPA Guidelines for Cancer Risk assessment (1986a). The new
assessment uses body weight raised to the 3/4-power as recommended by the EPA
Guidelines for Carcinogen Risk Assessment (U. S. EPA, 2005a).
• Different dose conversion factors were used in the conversion from ppm to mg/kg-d. The
new assessment applied strain-specific food consumption parameters (U.S. EPA, 1988a)
and body weights derived from growth curves in the study publication (NCI, 1978).
DDE —January, 2008
8-5
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• The linear multistage model was used for the 1988 IRIS assessment. The multistage (2)
model from BMDS Version 1.3.2 was used in the current assessment.
The linear multistage model derived a slope by fitting the dose-response data to a straight
line that passes through the zero point for the X and Y axises. The Guidelines for
Carcinogen Risk Assessment (U.S. EPA, 2005a) determine the slope factor by drawing a
straight line from the lower bound on the 10 percent response level (e.g., LED10) to zero.
Despite the difference in cancer risk assessment approaches listed above, the findings are
remarkably similar. The revised and legacy IRIS slope factors are 1.7 x 10"1 and 3.4 x 10"1
(mg/kg-day)"1, respectively. The revised and legacy IRIS HRL values are 0.2 |ig/L and 0.1 |ig/L,
respectively.
DDE — January, 2008 8-6
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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 (U.S. 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, U.S. EPA, 1997b). After review of and
response to comments, the final CCL was published on March 2, 1998 (63 FR 10273, U.S. EPA,
1998c).
On July 18, 2003, EPA announced final Regulatory Determinations for one microbe and
8 chemicals (68 FR 42897, U.S. EPA, 2003) after proposing those determinations on June 3,
2002 (67 FR 38222, U.S. EPA, 2002b). The remaining 41 chemicals and ten microbial agents
from the first CCL became CCL 2 and were published in the Federal Register on April 2, 2004
(69 FR 17406, U.S. EPA, 2004).
EPA proposed Regulatory Determinations for 11 chemicals from CCL2 on May 1, 2007
(72FR 24016) (U.S. EPA, 2007). Determinations for all 11 chemicals were negative based on a
lack of national occurrence at levels of health concern. The Agency is given the freedom to
determine that there is no need for a regulation if a chemical on the CCL fails to meet one of
three criteria established by the SDWA and described in section 9.1.1. After review of public
comments and submitted data, the negative determinations for the 11 contaminants have been
retained. Each contaminant will be considered in the development of future CCLs if there are
changes in health effects and/or occurrence.
9.1.1 Criteria for Regulatory Determination
These are the three 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.
• 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 the 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
DDE — January, 2008 9-1
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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 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 the 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 difference 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 judgment 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 DDE 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 of action, the critical effect(s), the dose-response for
critical effect(s), RfD for threshold effects, and the slope factor for non-threshold effects.
A full description of the health effects associated with exposure to DDE is presented in
Chapter 7 of this document and summarized below in Section 9.2.2. Chapter 8 and Section 9.2.3
present dose-response information.
DDE — January, 2008 9-2
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9.2.1 Health Criterion Conclusion
The available toxicological data indicate that DDE has the potential to cause adverse
health effects in humans and animals. DDE toxicity is manifested in a variety of effects
impacting the liver, nervous system, reproductive system, and possibly the immunological
system; however, toxicity varies depending on the species. There are limited human and animal
studies examining noncancer effects of chronic DDE exposure. The health reference level
(HRL; 0.2 |ig/L) for DDE is based on the occurrence of liver tumors in mice following chronic
exposures (NCI, 1978). There is no RfD for DDE; however, there is an RfD of 0.0005
mg/kg/day for DDT that is relevant to DDE as a derivative of DDT. Based on these
considerations, the evaluation of the first criterion for DDE is positive: DDE may have an
adverse effect on human health.
9.2.2 Hazard Characterization and Mode of Action Implications
DDE is not produced as a commercial product. This has limited the numbers of
conventional studies that have been performed to assess toxicological properties. Limited data
on DDE, mostly from the National Cancer Institute (NCI) bioassay, suggest that the liver is the
principle target organ in mammalian species. However, the NCI study did not evaluate a full
array of noncancer endpoints. Data on DDT identify effects on the nervous and endocrine
systems as adverse effects that might also be seen with DDE because it is one of DDT's principal
metabolites. The limited data for DDE suggest that any effects on the nervous system are less
severe than those observed with DDT. Endocrine effects from DDE are discussed in this section.
Based on animal studies DDE is likely to be carcinogenic to humans. This classification
is based on increases in the incidence of liver tumors, including carcinomas, in two strains of
mice and in hamsters after dietary exposure to DDE. Some epidemiological studies suggest a
possible association of the levels of DDE in serum with breast cancer. However, other studies
with similar methodologies do not show any association. DDE was mutagenic in mouse
lymphoma L5178Y and Chinese hamster V79 cells but negative in the Ames assay.
There are some indications that DDE has an adverse impact on the immune system
(Banerjee et al., 1996). Oral exposures to 22 mg/kg/day for six weeks suppressed serum
immunoglobin levels and antibody liters. Inhibition of leucocytes and macrophage migration
were observed at the cellular level. Considerable evidence exists that DDE can act as an
endocrine disrupter since it binds to the estrogen and androgen receptors. DDE has a stronger
affinity for the androgen receptor than for the estrogen receptor. It competes with testicular
hormones for the androgen receptor leading to receptor-related changes in gene expression
(Kelceetal., 1995).
EPA evaluated whether health information is available regarding the potential effects on
children and other sensitive populations. Children and adolescents may be sensitive populations
for exposure to DDE due to its endocrine disruption properties. Some data suggest that DDE can
delay puberty in males (ATSDR, 2003).
DDE — January, 2008 9-3
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9.2.3 Dose-Response Characterization and Implications in Risk Assessment
There is an RfD of 0.0005 mg/kg/day for the parent pesticide DDT based on an NOAEL
of 0.05 mg/kg/day from a dietary subchronic study (U.S. EPA, 1988b). In this study, liver
lesions were identified at an LOAEL of 0.25 mg/kg/day.
In the 1988 IRIS, EPA calculated an oral slope factor of 0.34 (mg/kg/day)4 for DDE. For
regulatory determination, EPA calculated an oral slope factor from the same data set (from the
1988 IRIS) using the EPA 2005 Cancer Guidelines (U.S. EPA, 2005a). The revised slope factor
is 1.67 x 10 ~l (mg/kg/day)"1 resulting in a one-in-a-million cancer-risk (HRL) of 0.2 |ig/L when
using exposure factors of 2 L/day drinking water ingestion and 70 kg body weight.
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 public health concern. 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 DDE 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 DDE, as well those that reported concentrations of DDE above an
estimated drinking-water HRL. For noncarcinogens, the estimated HRL 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 (i.e., the probability of 1 excess tumor in a
population of a million people). The HRLs are benchmark values that were used in evaluating
the occurrence data while the risk assessments for the contaminants were being developed.
The available monitoring data, including indications of whether or not the contaminant is
a national or a regional problem, are included in Chapter 4 of this document and summarized
below. Additional information on production, use, and fate is found in Chapters 2 and 3.
9.3.1 Occurrence Criterion Conclusion
DDE is a metabolite, degradation product, and impurity of DDT and is not commercially
used or produced. Although DDT was once widely used as a broad spectrum organochlorine
pesticide, it has been banned in the U.S. since January 1, 1973 and is not currently produced in
the U.S. Some areas of the world still produce and use DDT. DDE is persistent in the
DDE — January, 2008 9-4
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environment and mainly adheres to soil and sediments. Although DDE concentrations are
decreasing in the environment, monitoring data continue to show DDE in environmental media.
DDE has been analyzed in surface and ground water and in monitoring studies of
ambient water. At the MRL of 0.8 |ig/L, it was not detected in 3251 samples collected from 797
small systems under the UCMR 1. Of the 3077 large systems monitored, DDE was detected at
one large groundwater system, which represented 0.03% of large public water systems and
0.01% of the population served by them (approximately 18,000 people). The MRL of 0.8 |lg/L
was greater than the HRL and /^ the HRL. However, the MRL is equivalent to a risk of 4 x 10"6,
and thus, falls within the 10"4 to 10"6 risk that is the targeted in establishing a Maximum
Contaminant Level for a carcinogen. DDE also has been measured in sediment (maximum
concentration 31-440 |ig/kg) and in whole fish (maximum concentrations 450-7300 jig/kg).
Because the MRL was greater than the HRL, it is possible that DDE occurs in public
water systems at the HRL or 1A the HRL. However, this is considered unlikely due to the low
maximum levels observed in NAQWA ambient water monitoring where the detection limit and
the maximum concentration detected were below the HRL. Accordingly, the finding for the
second criterion is negative: DDE does not occur in public water systems at levels of concern.
9.3.2 Monitoring Data
Drinking Water
Occurrence data for DDE were collected through the UCMR 1 program. The first cycle
extended from 2001 to 2006. The minimum reporting level (MRL) is 0.8 |ig/L. A total of 797
small public water systems (590 ground water and 207 surface water) were tested with 3251
samples obtained. Among the small systems, DDE was not detected in any system. A total of
3077 large public water systems (1381 ground water and 1696 surface water) were tested with
30,546 samples obtained. Among the large systems, DDE was detected in a single ground water
system that represented 0.03% of large public water systems and 0.01% of the population served
by them (approximately 18,000 people). Because the minimum reporting level is greater than
half the Health Reference Level (^HRL or 0.1 |ig/L) and the full HRL (HRL or 0.2 |ig/L), it
cannot be stated how many samples may have contained DDE at the HRL or /^ the HRL.
Ambient Water
Occurrence data for DDE were collected with the NAWQA program from 1992 to 2001
(Cycle 1) in representative watersheds and aquifers across the country. Reporting limits varied
over the course of the cycle, but the level of detection did not exceed 0.006 |ig/L. In agricultural
areas, 1885 samples from 78 ambient surface water sites were collected and analyzed; the
detection frequency was 4.84%. Samples were also collected from 1443 ambient ground water
wells and analyzed; the detection frequency was 3.26%. The maximum concentration was 0.062
|ig/L in ambient surface water and 0.008 |ig/L in ambient ground water. In mixed land use
areas, 1021 samples from 47 ambient surface water sites were tested with a detection frequency
of 6.14% and 2716 samples from ambient ground water wells were tested with a detection
frequency of 2.65%. The maximum concentration was 0.009 |ig/L in ambient surface water and
0.006 |ig/L in ambient ground water. In undeveloped areas, 60 samples from 4 ambient surface
DDE — January, 2008 9-5
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water sites were tested with a detection frequency of 3.66% and 67 ambient ground water wells
were tested with a detection frequency of 7.46%. The maximum concentration was 0.002 |ig/L
in ambient surface water and in ambient ground water. In urban areas, 900 samples from 33
ambient surface water sites were tested with a detection frequency of 1.68% and 834 ambient
ground water wells were tested with a detection frequency of 3.96%. The maximum
concentration was 0.007 |ig/L in ambient surface water and 0.005 |ig/L in ambient ground water.
The median and 95th percentile concentrations were below the reporting limit and well below the
HRL and V* the HRL.
9.3.3 Use and Fate Data
DDE has no commercial use and is currently not commercially produced in the United
States. It is a degradation product and impurity of the once commercially produced pesticide,
DDT. DDT was once widely used as a broad spectrum organochlorine pesticide that controls
insects in agriculture and those that carry diseases such as malaria and typhus (Gianessi and
Puffer, 1992). Production of DDT in the United States was at its peak in 1962. On January 1,
1973, DDT use in the United States was banned. Analytical studies have revealed that DDE may
be a contaminant in technical grade insecticide dicofol (Risebrough et al., 1986). In addition,
DDE can be a product of the degradation of l,l,2,2-tetrachloro-2,2-bis(p-chlorophenyl)ethane,
another DDT-related impurity in dicofol (ATSDR, 2002).
DDE has high partition coefficients. Therefore, it is expected to adsorb strongly onto
organic matter such as soils, sediment, and suspended particulate matter. As a result of DDE's
strong binding to soil, it will remain mostly on the surface layers of soil (top 1.5 cm layer); small
amounts leach into the lower soil layers and groundwater (Callahan et al., 1979). Soils and
sediments are the environmental sink for DDE. DDE has a low water solubility (<0.12 mg/L).
DDE can bioaccumulate due to its high lipophilicity and long half-life, and may
biomagnify up the food chain. Callahan et al. (1979) reported the findings from a
terrestrial-aquatic microcosm experiment on the fate of 3.8 ppb DDE in water. The
bioconcentration factors (BCFs) were calculated to be 3.6xlO+4, 5.9xlO+4, 1.2xlO+4, and l.lxlO+4
for snail, mosquito larvae, fish, and algae in this experiment. NAWQA has measured DDE in
whole fish at levels ranging up to 7300 |ig/kg.
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 to levels 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 summarized in section 9.4.2 below.
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
DDE — January, 2008 9-6
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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 does regulation of the
contaminant in drinking water. In making the 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 below.
In making its regulatory determination, EPA also evaluated effects on potentially
sensitive populations, including the fetus, infants and children. Sensitive population
considerations are included in section 9.4.4.
9.4.1 Risk Criterion Conclusion
There are about 18,000 people (0.01% of the population served by monitored PWSs)
served by systems with DDE at the MRL (>0.8 |ig/L). It is not possible to estimate the
population exposed at either the HRL of /^ the HRL because both concentrations are less than
the MRL. DDE concentration in the air is low as shown by limited data from North Dakota
taken in 1993-94 in which air samples from two rural sites had DDE levels ranging from 6 to
200 pg/m3. Therefore, exposure via inhalation is considered negligible. DDE also has been
detected in food (at concentrations up to 0.1020 ppm) but levels in food materials have generally
declined in the years since the use of DDT was banned in the United States. Due to its high Kow,
it is selectively partitioned into fatty tissue. Because of this, DDE is found in breast milk, milk,
cheese, and fish. Estimated intakes of DDE in food during 1986 to 1991 were highest in infants
(0.0441 |ig/kg body weight/day). Intakes in adults were estimated to be between 0.0082 and
0.0119 |ig/kg body weight/day.
Although the risk of exposure to DDE from drinking water cannot be determined due to
the MRL exceeding the HRL, it would not exceed the estimated risk at the MRL, which is a risk
of four excess cancer cases in a million exposed persons. Based on the levels found in ambient
water it is unlikely that the levels of DDE in the public water system would exceed the HRL or
/^HRL. On the basis of these observations, the impact of regulating DDE concentrations in
drinking water on health risk reduction is likely to be small. Thus, the evaluation of third
criterion is negative: regulation of DDE in public water systems would not reduce the risk to the
population.
9.4.2 Exposed Population Estimates
There are no exposed population estimates at the HRL or /^ the HRL due to the fact that
the MRL was greater than the HRL. There was only one large groundwater system where DDE
was detected above the MRL, and the maximum concentrations of DDE in the ambient water
were all below the HRL and /^HRL. Therefore, it is likely that only a small portion of the
population is exposed to DDE in public water systems at either the HRL or
DDE — January, 2008 9-7
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9.4.3 Relative Source Contribution
Relative source contribution analysis compares the magnitude of exposure expected via
drinking water to the magnitude of exposure from intake of DDE in other media, such as food,
air, and soil. Exposure via air is negligible. Due to DDE's low water solubility, persistence, and
accumulation in lipids, food is likely to be the main source of exposure. DDT is still used in
other countries and the U.S. imports food from these countries. Therefore, DDE may still be a
concern in the food supply.
9.4.4 Sensitive Populations
The data from human studies are mostly epidemiological studies that compare the levels
of DDE in serum or adipose tissues to a variety of health outcomes. In many cases there are
conflicting study results even when there are strong similarities in methodologies. There is some
evidence that elevated serum levels in pregnant women are correlated to premature delivery and
lower infant birth weights. DDE may also have adverse effects on growth in children; these
results, however, are not conclusive. Breast feeding of infants exposes them to DDE stored in
their mothers adipose tissues as these stores are mobilized during lactation.
Animal studies suggest that males may be more susceptible than females to the endocrine
disrupting effects of DDE. DDE is an antiandrogen and effects on the development of male
sexual organs have been observed in a variety of rodent studies. The affinity of DDE for
androgen receptors is stronger than its affinity for estrogen receptors.
9.5 Regulatory Determination Decision
As stated in Section 9.1.1, a positive finding for all three criteria is required in order to
make a determination to regulate a contaminant. In the case of DDE, only the finding for the
criterion on health effects is positive. DDE may have an adverse effect on human health. Based
on the available drinking water monitoring data, DDE has been detected in one large
groundwater PWS above the MRL of 0.8 |lg/L, serving 17,670 people (0.01% of the population
served by large PWSs). DDE was detected at very low concentrations in ambient water. The
highest detection frequency was 7.46% in ambient ground water, and the highest maximum
concentration detected was 0.062 jig/L (which was detected in ambient surface water).
Commercial use of DDT, the parent compound of DDE, was phased out in the United States as
of January 1, 1973. Accordingly, it appears that DDE does not occur in public water systems
with a frequency and at levels of public health concern at the present time. DDE, however, is
still present in the sediment and soils, thereby leading to its presence in fish and other food
sources. Regulation of DDE in public water systems, however, does not appear to present a
meaningful opportunity for health risk reduction.
DDE — January, 2008 9-8
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DDE — January, 2008 10-25
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DDE — January, 2008 10-26
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APPENDIX A: Abbreviations and Acronyms
ATSDR
BCF
BMDS
BMI
CCL
DDA
ODD
DDE
DDMU
DDNU
DDOH
EMAP
f.w.
FR
gd
HDL
HED
Hg
HRL
IRIS
LDL
LOAEL
MCLG
MDL
mg/kg-day
MRL
mRNA
NAWQA
NCOD
NDWAC
NOAEL
NOPES
NPDWR
NPL
NTP
PCBs
ppb
ppm
ppt
PWSs
Agency for Toxic Substances and Disease Registry
bioconcentration factor
Benchmark Dose Software
body mass index
Contaminant Candidate List
2,2-bis(chlorophenyl)acetic acid
dichlorodiphenyldichloroethane
1,1 -dichloro-2,2-bis(p-chlorophenyl)ethylene
l-chloro-2,2-bis[p-chlorophenyl]ethylene
2,2-bis(chlorophenyl)acetonitrile
2,2-bis(chlorophenyl)ethanol
Environmental Monitoring and Trends Program
fat weight basis
Federal Register
gestation days
high density lipoproteins
human equivalent doses
mercury
Health Reference Level
Integrated Risk Information System
organic carbon/water partitioning coefficient
octanol-water partitioning coefficient
low density lipoproteins
lowest observed adverse effect level
Maximum Contaminant Level Goal
method detection limit
milligrams per kilogram per day
Minimum Reporting Level
messenger ribonucleic acid
National Water Quality Assessment
National Drinking Water Contaminant Occurrence Database
National Drinking Water Advisory Council
no observed adverse effect level
non-occupational exposure to pesticides
National Primary Drinking Water Regulation
National Priorities List
National Toxicology Program
polychlorinated biphenyls
l,l-dichloro-2,2-bis(p-chlorophenyl)ethylene
parts per billion
parts per million
parts per trillion
Public Water Systems
DDE —January, 2008
Appendix A-l
-------
RfD reference dose
RLs reporting levels
SDWA Safe Drinking Water Act
SF slope factor
SRBC sheep red blood cells
SVOCs semivolatile organic compounds
U.S. FDA United States Food and Drug Administration
U.S. EPA United States Environmental Protection Agency
UCMR 1 Unregulated Contaminant Monitoring Rule
VOCs volatile organic compounds
w.w. wet weight basis
WHO World Health Organization
DDE —January, 2008 Appendix A-2
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APPENDIX B: BMD Modeling Output
female mice
NCI dose
NCI food
consumption
(kg/day)
0 0.0061
148 0.0061
261 0.0061
NCI
NCI dose NCI fern NCI fern HED
terminal (mg/kg-
BW(kg) day)
0.037 0
0.031 29.12258
0.03 53.07
n
0
19
34
19 0
47 4.160369
48 7.581429
Model
Log-logistic
Quantal linear
Gamma
Multistage (2)
Weibull
Log-probit
Probit
Logistic
Quantal quadratic
ED10
12.42
5.06
10.69
8.09
10.69
13.23
13.72
14.30
14.64
LED
3.13
4.03
4.19
4.19
4.19
7.72
10.34
10.79
13.05
AIC
125
124
125
125
125
125
128
129
125
p-value
1.00
0.66
1.00
1.00
1.00
1.00
0.17
0.13
0.53
male mice
NCI dose
NCI food
NCI
NCI dose NCI male NCI
HED
consumption terminal (mg/kg- male n
(kg/day) BW(kg) day)
0 0.0064 0.037 0 0 19 0
148 0.0064 0.0345 27.45507 7 41 3.922153
261 0.0064 0.033 50.61818 17 477.231169
Model
Log-probit
Quantal linear
Quantal quadratic
Log-logistic
Gamma
Multistage (2)
Weibull
Probit
Logistic
mean
ED10
19.90
12.93
23.43
18.91
18.64
17.69
18.64
24.72
26.06
16.42
LED
15.71
9.39
19.96
8.20
9.58
9.58
9.58
19.35
20.55
12.55
AIC
101
101
102
103
103
103
103
104
105
p-value
1.00
0.84
0.74
1.00
1.00
1.00
0.36
0.31
DDE —January, 2008
Appendix B-l
-------
Multistage Model with 0.95 Confidence Level
T3
"o
<
O
•*=
O
CD
LL
0.8
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0
' ' ' M if'-t ' ' ' ' ' ' :
or /in i IVUIt'Sta9® j
tjML) Lower Dound :
^-^ \
.^ -
/^ \
x /^ 1
/^ // 1
^"
BMDL BMD i
0 10 20 30 40 50
dose
13:5703/232005
DDE —January, 2008
Appendix B-2
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