EPA/600/R-08/086F 

May 2010 

An Exposure Assessment of 

Polybrominated Diphenyl Ethers

National Center for Environmental Assessment

Office of Research and Development

U.S. Environmental Protection Agency 

Washington, DC 
ii

DISCLAIMER 

 This document has been reviewed in accordance with U.S. Environmental Protection 

Agency policy and approved for publication. Mention of trade names or commercial products 

does not constitute endorsement or recommendation for use. 

ABSTRACT 

 Polybrominated diphenyl ethers, PBDEs, are a class of brominated flame retardants that, 

like other persistent organic pollutants (POPs), have been found in humans, wildlife, and biota 

worldwide. Unlike other POPs, however, the key routes of human exposure are thought to be 

from their use in household consumer products, and their presence in house dust, and not from 

dietary routes. The exposure of Americans to PBDEs was systematically evaluated in this study. 

The production and lifecycle of the formulated PBDE products were examined. Literature on 

their fate and presence in the environment was reviewed. Exposure media data on brominated 

diphenyl ether (BDE) congeners were combined with estimates of adult, childhood, and infant 

intake factors to estimate a total intake of PBDEs for these receptors. The exposure pathways 

evaluated included food and water ingestion, inhalation, and ingestion of and dermal contact 

with house dust. For the adult intakes, a body burden of PBDEs was simulated using a simple 

pharmacokinetic model. The predicted body burdens were com- pared with representative adult 

profiles of PBDEs in blood and milk. The adult intake dose of total PBDEs was estimated to be 

7.1 ng/kg body weight-day. Comparable published values for total PBDE exposure average 3 to 

16 ng/kg-day. The estimated adult intake dose of 7.1 ng/kg-day was predicted to result in a body 

burden of 31.0 ng/g lipid weight (lwt). This compared to body burdens of 36.3 ng/g lwt found in 

blood and 44.1 ng/g lwt found in mother’s milk. Adult food intake contributes about 10% to 

total exposure and was estimated in this analysis to be about 0.6 ng/kg-day. Comparable 

published values for food intake were about 0.5 to 2.0 ng/kg-day. The largest source 

contributing to U.S. PBDE exposure was found to be house dust, estimated as 6.4 ng/kg-day, and 

contributing 90% of the overall estimated intakes. These adult intakes and body burdens were 

derived as central tendency estimates for the US general population. Intakes from housedust 

exposures (ingestion plus dermal contact) could range up to an order of magnitude higher than 

calculated for this central tendency considering the highest PBDE concentrations found in dust. 

The 95th percentile adult body burden of total PBDEs found in the 2003/2004 National Health 

and Nutritional Evaluation Survey (NHANES) was 291 ng/g lwt, about an order of magnitude 

higher than the central tendencies of 36.3 ng/g lwt in blood. Children’s estimated intakes were 

higher at 47.2 ng/kg/day for ages 1–5, 13.0 ng/kg/day for 6–11, and 8.3 ng/kg/day for 12–19. 

Infant intakes due to ingestion of mother’s milk were the highest at 141 ng/kg/day. Other 

exposures of interest examined in this report include fetal exposures, occupational exposures, 

and unusually high exposures in the general population. 

 

Preferred Citation: 

U.S. Environmental Protection Agency (EPA). (2010) An exposure assessment of polybrominated diphenyl ethers. 

National Center for Environmental Assessment, Washington, DC; EPA/600/R-08/086F. Available from the 

National Technical Information Service, Springfield, VA, and online at http://www.epa.gov/ncea. 
iii

CONTENTS

LIST OF TABLES ......................................................................................................................... vi

LIST OF FIGURES ..................................................................................................................... viii

ABBREVIATIONS AND ACRONYMS ...................................................................................... ix

PREFACE .................................................................................................................................... xiv

AUTHORS AND REVIEWERS .................................................................................................. xv

EXECUTIVE SUMMARY ........................................................................................................ xvii

1. INTRODUCTION ................................................................................................................ 1-1 

REFERENCES FOR CHAPTER 1 ............................................................................................. 1-4 

2. PRODUCTION, USE, AND LIFE CYCLE OF POLYBROMINATED DIPHENYL 

ETHERS ............................................................................................................................... 2-1 

2.1. INTRODUCTION ....................................................................................................... 2-1 

2.2. PRODUCTION ........................................................................................................... 2-2 

2.3. USES OF PBDEs ........................................................................................................ 2-6 

2.3.1. PBDE Content of Consumer Products ............................................................ 2-8 

2.4. GENERALIZED LIFE CYCLE OF PBDEs ............................................................. 2-11 

2.4.1. Production Releases ...................................................................................... 2-11 

2.4.2. Estimates of PBDE Releases to Indoor Air Based on Chamber Testing ...... 2-14 

2.4.3. Estimates of PBDE Releases to Air from Plastic Consumer Products .......... 2-17 

2.4.4. Estimates of the Mass Flow of PBDEs Contained in Electronic and 

Electrical Equipment (EEE) Waste in the United States ............................... 2-27 

2.4.5. Estimates of the Mass Loading of PBDEs Present in Sewage Treatment 

Plant Sewage Sludge and Effluent to Land and Water in the United 

States ............................................................................................................. 2-32 

2.5. SUMMARY OF THE GENERALIZED LIFE CYCLE ANALYSIS OF 

PBDES IN THE UNITED STATES ......................................................................... 2-35 

2.6. SUMMARY .............................................................................................................. 2-39 

REFERENCES FOR CHAPTER 2 ........................................................................................... 2-40 

3. ENVIRONMENTAL FATE OF POLYBROMINATED DIPHENYL ETHERS ............... 3-1 

3.1. INTRODUCTION ....................................................................................................... 3-1 

3.2. PHYSICAL/CHEMICAL PROPERTIES ................................................................... 3-1 

3.2.1. Water Solubility (WS) ..................................................................................... 3-1 

3.2.2. Octanol Water Partition Coefficient (Kow) ...................................................... 3-2 

3.2.3. Henry’s Law Constant (H) .............................................................................. 3-2 

3.2.4. Vapor Pressure (Vp) ......................................................................................... 3-7 

3.2.5. Octanol Air Partition Coefficient (Koa) ........................................................... 3-9 

3.2.6. Vapor Particle Partitioning in Air ................................................................. 3-14 

3.3. BIOACCUMULATION, BIOCONCENTRATION, AND 

BIOMAGNIFICATION OF PBDEs IN THE AQUATIC ENVIRONMENT .......... 3-16 
iv

CONTENTS (continued)

3.4. BIOTIC AND ABIOTIC DEBROMINATION AND TRANSFORMATION 

OF PBDEs ................................................................................................................. 3-17 

3.4.1. Microbial Degradation of PBDEs ................................................................. 3-19 

3.4.2. In Vivo Metabolic Debromination in Animals .............................................. 3-21 

3.4.2.1. Evidence for Debromination in Fish .............................................. 3-22 

3.4.2.2. Evidence for Debromination in Rats .............................................. 3-25 

3.4.2.3. Evidence for Debromination in Birds ............................................ 3-26 

3.4.2.4. Evidence for Debromination in House Cats ................................... 3-27 

3.4.3. Abiotic Degradation of PBDEs ..................................................................... 3-28 

3.4.3.1. Photodegradation of PBDEs ........................................................... 3-28 

3.4.3.2. Reaction with the Hydroxyl Radical .............................................. 3-37 

3.5. THERMAL DECOMPOSITION OF PBDE ............................................................. 3-37 

3.6. PATTERNS OF ENVIRONMENTAL FATE OF PBDE ......................................... 3-38 

3.6.1. Fate of PBDEs in Outdoor and Indoor Air .................................................... 3-39 

3.6.1.1. Fate of PBDEs in outdoor air ......................................................... 3-39 

3.6.1.2. Fate of PBDEs in Indoor Air .......................................................... 3-42 

3.6.2. Fate of PBDEs in Water ................................................................................ 3-44 

3.6.3. Fate of PBDEs in Soil ................................................................................... 3-45 

3.6.4. PBDEs in Sewage Treatment Plant Influent, Effluent, and Sludge .............. 3-47 

3.7. BIOACCUMLATION IN BIOTA ............................................................................ 3-52 

3.7.1. Bioaccumulation in the Aquatic Environment .............................................. 3-52 

3.7.2. Bioaccumulation in the Terrestrial Environment .......................................... 3-54 

3.7.2.1. Bioaccumulation in Birds ............................................................... 3-54 

3.7.2.2. Bioaccumulation in Terrestrial Mammals ...................................... 3-59 

3.7.2.3. Bioaccumulation in Insects ............................................................ 3-61 

3.8. ENVIRONMENTAL TIME TRENDS ..................................................................... 3-63 

3.8.1. Time Trends from Sediment Core Studies .................................................... 3-63 

3.8.2. Time Trends from Aquatic Wildlife Samples ............................................... 3-70 

3.9. CONCLUSIONS ....................................................................................................... 3-71 

REFERENCES FOR CHAPTER 3 ........................................................................................... 3-73 

4. ENVIRONMENTAL AND EXPOSURE MEDIA CONCENTRATIONS ......................... 4-1 

4.1. INTRODUCTION ....................................................................................................... 4-1 

4.2. WATER AND SEDIMENT CONCENTRATIONS ................................................... 4-2 

4.3. SURFACE SOIL CONCENTRATIONS .................................................................... 4-8 

4.4. DUST CONCENTRATIONS ................................................................................... 4-10 

4.4.1. House Dust Concentrations ........................................................................... 4-10 

4.4.2. Car and Airplane Dust Concentrations .......................................................... 4-27 

4.5. AIR CONCENTRATIONS ....................................................................................... 4-29 

4.5.1. Outdoor Air Concentrations .......................................................................... 4-29 

4.5.2. Indoor Air Concentrations and Simultaneous Indoor/Outdoor 

Monitoring ..................................................................................................... 4-50 
v

CONTENTS (continued)

4.6. FISH CONCENTRATIONS ..................................................................................... 4-56 

4.6.1. Farmed Fish Concentrations .......................................................................... 4-56 

4.6.2. Freshwater and Marine Fish Concentrations ................................................. 4-67 

4.6.3. Fish from the Retail Marketplace .................................................................. 4-73 

4.6.4. Observations from Fish Data ......................................................................... 4-76 

4.7. FOOD CONCENTRATIONS ................................................................................... 4-77 

4.8. ASSIGNING EXPOSURE MEDIA CONCENTRATIONS FOR EXPOSURE 

ASSESSMENT PURPOSES ..................................................................................... 4-93 

REFERENCES FOR CHAPTER ........................................................................................... 4-101 

5. HUMAN EXPOSURE .......................................................................................................... 5-1 

5.1. INTRODUCTION ....................................................................................................... 5-1 

5.2. BODY BURDEN DATA ............................................................................................ 5-1 

5.2.1. Blood Data ....................................................................................................... 5-2 

5.2.2. Human Milk Data .......................................................................................... 5-23 

5.2.3. Adipose Tissue Data ...................................................................................... 5-37 

5.2.4. Selection of Representative Body Burden Profiles ....................................... 5-38 

5.3. STUDIES ON INTAKE OR EXPOSURE DOSE .................................................... 5-41 

5.4. ESTIMATES OF BACKGROUND INTAKES OF PBDEs FOR ADULTS ........... 5-56 

5.5. CONVERTING ADULT INTAKE DOSE TO BODY BURDEN ........................... 5-61 

5.6. EXPOSURE OF SPECIAL POPULATIONS OF INTEREST TO PBDEs .............. 5-66 

5.6.1. Body Burden Data to Characterize Exposure to the Fetus ............................ 5-66 

5.6.2. Impacts to Infants from Consumption of Breast Milk .................................. 5-73 

5.6.3. Childhood Intakes .......................................................................................... 5-79 

5.6.4. Intake Estimates Derived for EPA’s VCCEP and by the NAS ..................... 5-80 

5.6.5. Body Burden Data to Characterize Occupational Exposures ........................ 5-83 

5.6.6. Elevated Exposures at the High End of the General Adult Population ......... 5-86 

5.7. UNCERTAINTY AND VARIABILITY IN ESTIMATING INTAKE DOSE 

AND CONVERTING THAT DOSE TO A BODY BURDEN ................................ 5-91 

5.7.1. Uncertainties with Estimates of Dust Intakes of PBDE ................................ 5-91 

5.7.2. Uncertainties and Variabilities with Other Pathways .................................... 5-94 

5.7.3. Uncertainties with Pharmacokinetic Modeling ............................................. 5-96 

5.8. OVERALL FINDINGS OF EXPOSURE OF AMERICANS TO PBDEs ............... 5-99 

REFERENCES FOR CHAPTER 5 ......................................................................................... 5-102 
vi

LIST OF TABLES 

1-1. BDE congener numbers and chemical composition of the most commonly studied 

BDE congeners ................................................................................................................ 1-2 

2-1. Approximate BDE homologue- and congener-specific content of PBDE 

commercial formulations ................................................................................................. 2-3 

2-2. Mean concentration (mg/g) of penta- and octaBDE in various flame retarded 

polymers ........................................................................................................................... 2-9 

2-3. Mean concentrations of PBDEs (mg/kg) found in EEE waste material at a 

recycling plant in Switzerland ......................................................................................... 2-9 

2-4. BDE congener concentrations (mg/kg) present in EEE waste components 

sampled at a recycling facility in 2002 .......................................................................... 2-10 

2-5. Estimated environmental releases (kg) of BDE congeners from U.S. decaBDE 

production and manufacturing facilities in 2007 ........................................................... 2-14 

2-6. Calculated PBDE emission factors and estimated PBDE emissions from plastic 

products in the United States for 2001, derived using two approaches: vapor 

pressure approach; log Koa approach ............................................................................. 2-22 

2-7. Estimated BDE congener emissions (kg/year) to air in the United States from 

incineration of EEE waste and estimated BDE congeners in EEE waste that is 

landfilled and recycled ................................................................................................... 2-27 

3-1. Estimated water solubility values for PBDEs .................................................................. 3-3 

3-2. Estimated octanol water partition coefficients (log Kow) values for PBDEs ................... 3-5 

3-3. Estimated Henry’s Law constants (H) for PBDEs ........................................................... 3-8 

3-4. Estimated solid phase vapor pressures (PS) and subcooled liquid vapor pressures 

(PL) of some PBDEs ...................................................................................................... 3-10 

3-5. Estimated octanol air partition coefficients (log Koa) of PBDEs ................................... 3-13 

3-6. Calculated theoretical vapor particle partitioning of PBDE congeners in ambient 

air at 25°C ...................................................................................................................... 3-15 

3-7. Estimated BAF and BMF values for various aquatic species ........................................ 3-18 

3-8. The BDE congener distributions in influent, sludges, and final effluent from 

various surveys of sewage treatment plants in various countries .................................. 3-50 
vii

LIST OF TABLES (continued)

4-1. Congener-specific concentrations of PBDEs in house dust in the United States .......... 4-11 

4-2. Outdoor and indoor congener-specific air concentrations of PBDEs in the 

United States .................................................................................................................. 4-30 

4-3. Congener-specific concentrations of PBDEs for fish caught in the United States ........ 4-57 

4-4. Congener-specific concentrations of PBDEs in food originating from the United 

States .............................................................................................................................. 4-78 

4-5. Exposure media concentrations ..................................................................................... 4-97 

5-1. Blood concentrations of PBDE congeners in Americans ................................................ 5-3 

5-2. Breast milk concentrations of PBDE congeners in Americans ..................................... 5-24 

5-3. Representative body burden levels of PBDEs in Americans ......................................... 5-40 

5-4. Estimates of general population intake or exposure dose of total PBDEs 

provided in the literature ................................................................................................ 5-42 

5-5. Exposure pathways and factors for the PBDE intake dose estimate ............................. 5-57 

5-6. Congener-specific and total adult intake estimates of PBDEs ....................................... 5-59 

5-7. Pharmacokinetic parameters and predicted concentrations of BDEs in adults 

compared with measurements in blood and milk .......................................................... 5-64 

5-8. Comparison of median congener concentrations from NHANES and median 

congener concentrations from a large study of umbilical cord blood 

measurements of PBDEs................................................................................................ 5-68 

5-9. Pharmacokinetic parameters for modeling the body burden impacts to infants 

via breast feeding, and then to children from food and household exposures ............... 5-75 
viii

LIST OF FIGURES 

2-1. General structure of polybrominated diphenyl ethers ...................................................... 2-2 

2-2. Idealized life cycle of polybrominated diphenyl ethers if all information is 

available to complete the analysis .................................................................................. 2-12 

2-3. TRI data showing environmental releases (kg) of decaBDE from production and 

manufacturing facilities in 2007, and total environmental releases from 1998 to 

2007................................................................................................................................ 2-13 

2-4. Volatilization of BDE 47 from PBDE-treated consumer products in the UK ............... 2-19 

2-5. Summary of a generalized annual life cycle of PBDE in the United States: 

production, use, disposal, recycling, and sewage treatment .......................................... 2-36 

3-1. Time trends of deposition flux of PBDEs (blue solid diamonds) and PCBs 

(red open diamonds) to the sediments in each of the Great Lake .................................. 3-65 

5-1. Approach for characterizing exposure to PBDEs in this report ....................................... 5-2 

5-2. Modeled infant and childhood body burdens of PBDEs ............................................... 5-76 

5-3. The fraction of the maximum concentrations of PCDD/F TEQ and BDE 47 

concentrations found at various percentiles within NHANES surveys of these 

two contaminants in adults ............................................................................................. 5-87 
ix

ABBREVIATIONS AND ACRONYMS

ABS acrylonitrile butadiene styrene 

BAF bioaccumulation factor 

BCF bioconcentration factor

BDE brominated diphenyl ether; used for congener designation as in 

BDE 47 

BFR brominated flame retardant

BMF biomagnifications factor 

BMP biomagnifications potential

BW body weight 

CARB California Air Resources Board

CDC Centers for Disease Control and Prevention 

CDD/Fs chlorinated dibenzodioxins and dibenzofurans 

CHDS Child Health and Development Studies 

CPSC Consumer Products Safety Commission 

CSFII Continuing Survey of Intakes by Individuals 

DE diphenyl ether 

decaBDE since there are only 10 available spots for bromines, this is 

synonymous with the BDE congener with 10 bromines, which is 

BDE 209. decaBDE is also the name of the formulation dominated 

(>97%) by BDE 209 

diBDE the homologue group of BDE congeners with two bromines 

DL detection limit

DBDPE decabromodiphenyl ethane 

DBDPO [

14C]decabromodiphenyl oxide 

DSBG Dead Sea Bromine Group

dwt dry weight 

EEE electronics and electrical equipment

EF emission factor

EP epoxy resin 

EPA Environmental Protection Agency

EU European Union 

EWG Environmental Working Group 
x

ABBREVIATIONS AND ACRONYMS (continued)

FPUF flexible polyurethane foam 

FSA Food Standards Agency of the UK 

GC/GC-IDTOFMS gas chromatography and isotope dilution time-of-flight mass 

spectrometry 

GC-IDHRMS gas chromatography and isotope dilute high resolution mass 

spectrometry

GC/MS Gas Chromatography / Mass Spectrometry

H Henry’s Constant 

HBCD hexabromocyclododecane 

heptaBDE the homologue group of BDE congeners with seven bromines 

hexaBDE the homologue group of BDE congeners with six bromines 

Hi-Fi high fidelity 

HIPS impact polystyrene 

HRGC high resolution gas chromatography 

HRMS high resolution mass spectrometry

IT information technology 

Koa octanol air partition coefficient

Kow octanol water partition coefficient

LOQ limit of quantification

lwt lipid weight 

MSW, MSWLF municipal solid waster, MSW landfill

MoBPXDD/Fs monobromo-polychlorinated dibenzo-p-dioxins and dibenzofurans 

monoBDE the homologue group of BDE congeners with one bromine 

NA not analyzed, or not available (as in data not available)

NAS National Academies of Science

ND not detected 

NEW Northwest Environment Watch 

NHANES National Health and Nutritional Examination Survey 

nonaBDE the homologue group of BDE congeners with nine bromines 

NSSS National Sewage Sludge Survey 
xi

ABBREVIATIONS AND ACRONYMS (continued)

octaBDE the homologue group of BDE congeners with eight bromines; this 

also refers to a commercial formulation of PBDEs dominated by 

hepta- and octaBDEs

PBDD/Fs polybrominated dibenzo-p-dioxins and dibenzofurans

PBDE polybrominated diphenyl ether 

PBT persistent and bioaccumulative toxic; in the context of describing 

types of plastics to which PBDEs are added, it also could be an 

abbreviation for polybutylene terephthalate 

PC personal computer

PCB polychlorinated biphenyl 

PCDD/Fs polychlorinated dibenzo-p-dioxins and dibenzofurans

PE polyethylene 

PERC perchloroethylene 

pentaBDE the homologue group of BDE congeners with five bromines; this 

also refers to a commercial formulation of PBDEs dominated by 

penta BDE congeners

PK pharmacokinetic 

POP persistent organic pollutants 

PP polypropylene 

PS solid phase vapor pressure in Pa 

PS-1 refers to the high volume air sampler 

PUF polyurethane foam 

PUR-H polyurethane hard foam 

RCRA Resource Conservation and Recovery Act 

RD reductive dehalogenase 

RfD reference dose

SML sea-surface microlayer

SD standard deviation

SRM Standard Reference Material

SSW subsurface seawater

STP sewage treatment plant
xii

ABBREVIATIONS AND ACRONYMS (continued)

Σ9BDEs summation of the concentrations of 9 key BDE congeners; other 

numerical subscripts would indicate a different number of 

congeners 

TCDD 2,3,7,8-tetrachlorodibenzo-p-dioxin

TCE trichloroethylene 

TEQ toxic equivalent; used in the context of dioxin, where TEQ is the 

concentration of a mixture of dioxin-like compounds toxically 

equivalent to 2,3,7,8-TCDD

tetraBDE the homologue group of BDE congeners with four bromine 

TRI Toxics Release Inventory

triBDE the homologue group of BDE congeners with three bromines 

TV television

UC University of California

UNEP United Nations Environmental Program 

UPE unsaturated polyesters

U.S. United States

UK United Kingdom 

UV ultraviolet 

VCCEP Voluntary Children’s Chemical Evaluation Program 

VCR video cassette recorder

Vp vapor pressure 

WHO World Health Organization 

WS water solubility

WWF World Wildlife Fund 

wwt wet weight

WWTP waste water treatment plant

WTC World Trade Center

XAD As in XAD resin, a highly absorbent resin used in continuous 

sampling of organic materials, especially for monitoring of 

pollutants in gas streams.

XRF X-ray fluorescence
xiii

ABBREVIATIONS AND ACRONYMS (continued)

Units of Measurement

atm atmospheres

C centigrade

cm centimeter

g gram 

h, hr hour 

kg kilogram 

km kilometer

L liter

m meter (used mostly as m3, as air volume, but also as a length 

measure)

mg milligram

μg microgram

μL microliter

μm micrometer (also micron)

μM micro molar

mL millileter

mm millimeter

MMT million metric tons

MT metric tons

mol mole, in chemistry

ng nanogram 

Pa Pascals

pg picogram

pmol picomole 

s seconds 

yr year
xiv

PREFACE

The U.S. Environmental Protection Agency (EPA) formed a working group composed of 

individuals from several program offices including the Offices of Pesticides, Prevention, and 

Toxic Substances, the Office of Water, and the Office of Research and Development, Office of 

Policy, Economics and Innovation, to study production, use, alternatives, environmental fate, 

exposure, and health effects of polybrominated diphenyl ethers (PBDEs). This working group 

issued a project plan in 2006 that outlined projects in these areas, and updated that plan in 

December 2008. EPA reports regularly on progress in completing the activities identified in the 

project plan, with the most recent status report issued in March 2008 and a planned update for 

2010. The Web site that describes this working group, including the project plan and status 

reports, is http://www.epa.gov/oppt/pbde. This document addresses the exposure assessment 

needs identified in that project plan. It provides a comprehensive assessment of the exposure of 

Americans to this class of persistent organic pollutants. Individual chapters in this document 

address the production, use, and life cycle of PBDEs; environmental fate; environmental levels;

and human exposure. 
xv

AUTHORS AND REVIEWERS 

The National Center for Environmental Assessment (NCEA), Office of Research and 

Development was responsible for the preparation of this document. 

AUTHORS

Matthew Lorber 

Exposure Analysis and Risk Characterization Group 

National Center for Environmental Assessment

U.S. Environmental Protection Agency 

Washington, DC 20460 

David Cleverly 

Exposure Analysis and Risk Characterization Group 

National Center for Environmental Assessment 

U.S. Environmental Protection Agency 

Washington, DC 20460 

INTERNAL EPA REVIEWERS

Linda Birnbaum 

Jeffrey Frithsen

Office of Research and Development

Daniel Axelrad

Greg Miller

Keeve Nachman

Office of Policy, Economics and Innovation 

Bob Boethling 

Lynn Delpire 

Tala Henry

Office of Pollution, Prevention and Toxic Substances 

EXTERNAL PEER REVIEWERS

Robert Hale 

Virginia Institute of Marine Sciences

S310 Chesapeake Bay Hall

Gloucester Point, VA 23062 
xvi

AUTHORS AND REVIEWERS (continued)

Stuart Harrad 

University of Birmingham 

Public Health Building, Room 209A 

Birmingham, West Midlands B15 2TT 

United Kingdom 

John Jake Ryan

Health Canada―Retired

Ottawa, ON K1V 9C6

Canada 

Daniele Staskal 

ToxStrategies

3420 Executive Center Drive 

Suite 114 

Austin, TX 78731 

Thomas Webster 

Boston University School of Public Health 

Department of Environmental Health

715 Albany Street, Talbot 4W 

Boston, MA 02118 

ACKNOWLEDGMENTS

The authors thank the staff of ECFlex, Inc. and IntelliTech Systems, Inc. for their 

editorial and word processing support. 
xvii

EXECUTIVE SUMMARY 

Polybrominated diphenyl ethers, PBDEs, are a class of brominated flame retardants 

intentionally manufactured to retard the combustibility of treated materials. When fire occurs, the 

PBDE formulations utilize vapor phase chemical reactions that interfere with the combustion 

process, thus delaying ignition and inhibiting the spread of fire. These characteristics have 

promoted the widespread use of PBDEs in textiles, flexible polyurethane foams used in 

upholstery stuffing for furniture and car seats, electronic components, electrical components, and 

plastics used in the casings of televisions, personal computers, and other electronic equipment. 

PBDEs have a common structure of a brominated diphenyl ether molecule that may have 

anywhere from 1 to 10 bromine atoms attached. Depending on the location and number of 

bromine atoms, there are 209 possible PBDE compounds; each are termed congener and are 

assigned a specific brominated diphenyl ether (BDE) number. PBDEs have been marketed in 

three primary formulations: (1) the “penta” formulation, commercially known as DE-71 and 

Bromkal 70-5DE; (2) the “octa” formulation, DE-79; and (3) the “deca” formulation, DE-83R or 

Saytex 102E. The formulations differ in their composition of BDE congeners. The dominant 

congeners in pentaBDE (percent weight basis in parenthesis) are BDE 99 (35−50%), BDE 47 

(25−37%), BDE 100 (6−10%), BDE 153 (3−5%), and BDE 154 (2−4%). The octa formulation 

is composed of BDE 183 (40%), BDE 197 (21%), BDE 203 (5−35%), BDE 196 (8%), BDE 208 

(10%), BDE 207 (7%), BDE 153 (5−10%), and BDE 154 (1−5%). The deca formulation is 

dominated by BDE 209 (97.5%), with the remainder being BDE 206 (2.2%), BDE 207 (0.24%), 

and BDE 208 (0.06%). 

Studies have been conducted in laboratory animals to gain a better understanding of the 

potential health risks of PBDEs. These studies have suggested potential concerns about liver 

toxicity, thyroid toxicity, developmental and reproductive toxicity, and developmental 

neurotoxicity. These findings raise particular concerns about the potential risks to children. The 

carcinogenic potential of some PBDEs have been studied. In a review of toxicological studies as 

part of a toxicological review for EPA’s Integrated Risk Information System (2008), EPA has 

found that the data for decabromodiphenyl ether (the commercial formulation composed mostly 

of BDE 209) support a finding of “suggestive evidence of carcinogenic potential” according to 

EPA’s Guidelines for Carcinogen Risk Assessment. For BDE congeners BDE 99, BDE 153, and 
xviii

BDE 47, EPA has found that the data support a finding of “inadequate information to assess 

carcinogenic potential.”

Production, Use and Release: 

Approximately 56,418 metric tons (MT) of PBDEs were produced worldwide in 2003, 

the latest reporting year, with between 40,000 and 67,000 MT/year produced between 1999 and 

2002. In 2001, decaBDE accounted for 83% of total worldwide production, followed by 

pentaBDE (11%), and octaBDE (6%). Approximately 95% of the global production of 

pentaBDE, 40% of octaBDE, and 44% of decaBDE were consumed in the Americas (North, 

South, and Central America). The penta and octa formulations were voluntarily withdrawn from 

the U.S. marketplace by their manufacturers at the end of 2004, leaving only the deca 

formulation currently being marketed for use in commercial products in the United States. The 

penta and octa formulations were banned in Europe, thus leaving the deca formulation as the 

only currently used formulation in Europe, too. However, in December of 2009, the EPA 

announced the phase out of decaBDE, with production, importation, and sales of decaBDE for 

most uses in the United States to end by 12/31/2012, and all uses to end by 12/31/2013. 

A limited life cycle analysis provides some indication of the magnitude of releases of 

PBDEs to the outside and inside environments, based on estimates of releases from PBDE 

production processes, from volatilization of PBDEs in products while in use, and from disposal 

of products containing PBDEs. Figure E-1 is a summary of a generalized annual life cycle of 

PBDE in the United States: production, use, disposal, recycling, and sewage treatment. The 

Toxics Release Inventory (TRI) reports a total release of 32.2 MT of decaBDE to the air, land, 

and water in 2007. According to TRI, total environmental releases peaked in 1999, with a 

release of 53.9 MT, and stayed at similar levels through 2002. There was a drop in releases in 

2003 to 36.3 MT, followed by an increase in 2004 to 44.8 MT, and then declines in 2005, 2006, 

and 2007. 

Much of the manufactured PBDEs have been used to make plastic consumer products 

flame retardant. A mathematical model was used to estimate the amount of PBDEs that may 

volatilize from the treated product to the air (mostly indoor air) during product use in the United 

States. It was estimated that 133.5 kg/year of BDE 47; 24.1 kg/year of BDE 99; 8.5 kg/year of 

BDE 209; 7.8 kg/year of BDE 28; and 3 kg/year each of BDE 66, BDE 153, and BDE 183 
xix

Figure E-1. Summary of a generalized annual life cycle of PBDE in the 

United States: production, use, disposal, recycling, and sewage treatment.

may volatilize to the air using this model. These estimates can, at best, be considered initial 

estimates that are unverifiable. However, they are consistent with chamber studies in that the 

greater the vapor pressure, and the lower the octanol air partition coefficient, the more likely the 

BDE congener will volatilize from the plastic product into the air. In this regard, BDE 47 has the 

greatest tendency to volatilize from plastic. Although uncertain, the estimates are judged to have 

provided an order of magnitude estimate of the volatile release of these flame retardants from 

plastic products in use. Another release mechanism noted is the flaking off, crumbling, and/or 

general physical removal of PBDEs from treated foam products. Although such a loss 

mechanism may be significant, there is no way to quantify this avenue of loss, or even 

qualitatively compare it to other loss mechanisms. 

Releases can occur from the disposal of electronic and electrical equipment (EEE) waste. 

In the United States, an estimated 2.4 million metric tons (MMT) of EEE waste was generated in 

2005. Of this amount, 0.3 MMT was recovered for recycling, 0.42 MMT was incinerated in 
xx

municipal waste combustors, and 1.68 MMT was buried in municipal solid waste (MSW)

landfills. Extrapolating to U.S. from limited European studies suggests that the EEE waste

contains about 1,224 MT decaBDE and 1,272 MT of octaBDE. However, U.S. EEE waste has 

yet to be adequately characterized for PBDE content. Losses of PBDEs in recycled products 

were not estimated in the life cycle analysis. The landfilling of EEE waste was assumed not to 

result in environmental releases of PBDEs, because of the careful conditions (liners, treated 

leachate) required by law for MSW landfills in the United States. However, the life cycle 

analysis suggests that MSW landfills are significant environmental reservoirs for PBDEs. It is 

possible that 1,747 MT PBDEs were disposed of in U.S. landfills in 2005 alone. Based on 

measured concentrations of decaBDE in EEE products and the destruction efficiency from MSW 

incinerators with a high degree of pollution control, releases of BDE 209 totaled 44.9 kg/year. 

Sewage treatment plant operations also result in environmental releases of PBDEs. The 

untreated sewage entering the treatment plant contains significant amounts of PBDEs. Both 

sewage sludge from the treatment plant processes as well as the treated sewage treatment plant 

effluent contain appreciable amounts of PBDEs. The annual land application of sewage sludge 

as a soil amendment practice results in 0.26, 2.69, 2.72, and 8.29 MT/year of BDE 153, BDE 47, 

BDE 153, and BDE 209, respectively, being applied to land in the United States. The life cycle 

analysis estimates that 37, 46.5, 82.1, 498, and 531 kg/year of BDE 154, BDE 153, BDE 209, 

BDE 47, and BDE 99, respectively, are discharged annually into surface waters in the United 

States from sewage treatment plant effluent.

Environmental Fate 

PBDEs are ubiquitous environmental contaminants. Widespread dispersion of BDE 

congeners in the environment is governed by their respective physical and chemical properties. 

The atmosphere is the primary transport media, and soils and sediments are environmental sinks. 

Transport can occur over relatively long distances, >1,000 km. Evidence for this comes from the 

presence of PBDEs in the polar environments, and in the tissues of deep ocean-dwelling whales 

and other marine mammals who spend a significant portion of their lives far from anthropogenic 

sources. PBDEs are lipophilic and hydrophobic compounds and readily bioaccumulate into 

terrestrial and aquatic food webs. This tendency has resulted in extensive accumulation of 

PBDEs in a wide variety of birds, fish, insects, and aquatic and terrestrial mammals reflective of 
xxi

dietary exposures through the food chain. In most fish (ocean and freshwater) and marine 

mammals, BDE 47 is the major congener contributing >30% to total body burden of PBDEs. 

The congener distribution in tissues of fish usually follows the order BDE 47 > BDE 99 

> BDE 100 > BDE 154 > BDE 153 > BDE 49 > BDE 28. In the few studies where BDE 209 

was measured in fish tissue, BDE 209 has ranged from <1 up to 88% of total PBDE body 

burden, depending on the dietary source of BDE 209. The PBDE body burden of birds is highly 

influenced by diet. For example, in a study of Peregrine Falcon eggs, the rank order by PBDE 

concentration in eggs (high to low) was: BDE 153 > BDE 99 > BDE 100 > BDE 154 > BDE 209 

> BDE 183 > BDE 197 > BDE 47 > BDE 207. BDE 47 was a minor constituent in Peregrine 

Falcon eggs (on average only 4.4%). The Peregrine Falcon consumes a diet consisting of 

medium sized birds such as doves, waterfowl, songbirds, waders, and pigeons. The Cormorant 

nests in swampy areas or near large bodies of water such as rivers and lakes, and its diet consists 

mainly of fish. BDE 47 contributed approximately 42% of the total body burden of PBDEs in 

the Cormorant. In the Common Buzzard, the most prevalent congeners detected were BDE 153 

(29% of total), BDE 99 (23%), and BDE 47 (22%). The Common Buzzard preys mainly on 

small mammals (mice, voles, rabbits, squirrels, rats, and moles). The blackbird, on the other 

hand, feeds principally on seeds and insects, and BDE 47 was the only congener detected in the 

adipose tissue of this species. Terrestrial mammals also tend to accumulate brominated PBDEs 

in relation to their diet. For example, the red fox is a top terrestrial predator that mainly 

consumes voles, rabbits, squirrels, and mice as prey, and BDE 209 dominated the total body 

burden. Maritime fish-eating bears displayed a BDE congener pattern dominated by BDE 47 

followed by BDE 209 > BDE 99 > BDE 100 > BDE 153. In the interior meat-eating bears, the 

tissues were dominated by BDE 209 followed by BDE 206 > BDE 47 > BDE 207 > BDE 208. 

Once released into the air, PBDEs partition between the vapor and particle phases in the 

atmosphere in accordance with their respective vapor pressures. Lower-brominated PBDEs 

primarily exist in the vapor phase, while higher-brominated congeners are primarily adsorbed to 

atmospheric particles. Photolysis (degradation by sunlight) in air is an important atmospheric 

removal process for BDE 47 and BDE 99. In contrast, the photolysis of BDE 209 in air is a 

minor atmospheric removal process. Atmospheric wet and dry surface deposition of BDE 209 is 

the most important atmospheric removal pathway. This is because BDE 209 predominantly 

exists bound to particles in air. Evidence suggests that BDE 209 can be degraded by ultraviolet 
xxii

light (i.e., photolysis) to form lower-brominated BDEs, and this may be an important degradation 

pathway in the environment. 

Several studies have shown that the higher-brominated BDE congeners can undergo 

biotic debromination. Several recent studies have provided evidence of microbial-mediated 

reductive debromination of decaBDE and octaBDE under laboratory conditions. For example, in 

one study, Sulfurospirillum multivorans bacterium incubated with decaBDE has induced 

reductive debromination of decaBDE in vitro to yield octa- and heptaBDE after a contact time of 

2 months. The octa- and heptaBDE did not further debrominate in the presence of the 

microbe―even after 1 year. The researcher of this study concluded that the microorganism was 

specific to the degradation of decaBDE and is incapable of debrominating lower-brominated

PBDE compounds. In another study, BDE 209 was debrominated in vitro to yield BDE 206, 

BDE 207, and BDE 208 by contact with anaerobic mesophilic microorganisms indigenous to raw 

sewage sludge (microbial species not identified). Methane was formed as a product of microbial 

respiration, and the amount of BDE 209 decreased by 30% within 238 days. 

Several studies provide evidence that in vivo metabolic debromination occurs in fish, 

birds, and mammals. Laboratory studies of rainbow trout, lake trout, and carp, involving fish 

food spiked with pure BDE 209, have clearly shown accumulation of lower-brominated BDE 

congeners not initially present in the feed. This evidence is suggestive of metabolic synthesis of 

lower-brominated congeners through debromination of BDE 209. Debromination in fish has 

also been observed in the wild. In a study at the outfall to a sewage treatment plant, sunfish 

bioaccumulated BDE 209. It was observed that the sunfish tissues contained BDE congeners not 

present in the sewage treatment plant effluent or in the surrounding aquatic environment. These 

were two octa- (BDE 201 and BDE 202) and three hepta- (BDE 188, BDE 184, and BDE 179) 

congeners. The investigating scientists proposed that the presence of these congeners in the 

sunfish was supportive of in vivo synthesis from the metabolic debromination from BDE 209. 

In another study, Sprague-Dawley rats were fed a commercial formulation of decaBDE. 

Evidence of metabolic debromination of BDE 209 to lower congeners was observed from an 

apparent 160% increase in the tissue concentration of BDE 197, BDE 201, and BDE 207 in the 

sacrificed rats as compared to levels in the feed. Additional evidence for in vivo metabolic 

debromination was found for chickens, starlings, and even house cats. 
xxiii

The study on house cats involved measuring the PBDE congener profile in cat fat and 

then also in the serum of house cats consuming dry food only, canned wet food only, and a 

combination of dry and wet cat food. The contamination of PBDEs in dry cat food reflected the 

congener profile of decaBDE, with BDE 209 representing 83−93% of total PBDE present in the 

food. Because of the high content of BDE 209 in dry cat food, BDE 209 dominated serum in 

cats only consuming the dry food. BDE 209 accounted for 4.2%, 21%, and 30% of serum PBDE 

levels in house cats consuming canned, mixed, and dry food, respectively. BDE 207 was 

consistently present in serum in significant concentrations of the dry food eaters as compared to 

the consumers of the other food types. BDE 207 accounted for 4.5%, 9.8%, and 17% of the 

PBDE levels detected in cats consuming canned, mixed, and dry food, respectively. BDE 207 

was present in the dry food-eaters at approximately 50% of the total concentration of BDE 209, 

which is uncharacteristic of the decaBDE congener profile (BDE 207 is approximately 1% of 

BDE congeners present in decaBDE), and was not the pattern observed in the wet food-eaters. 

Moreover, the ratio of BDE 207 to BDE 209 in cat serum was relatively constant in all dry 

food-eaters. The authors regarded these data as possible evidence for the metabolic 

debromination of BDE 209 to form BDE 207. 

 

Human Exposure and Biomonitoring

PBDEs have captured the attention of scientists and policymakers because levels in the 

environment and humans have increased rapidly since these chemicals came into use in the 

1960s and 1970s. Environmental time-trends can be observed from lake sediment core studies 

and archived animal tissue samples. The sediment cores show predominance for, and a stark rise 

in, BDE 209, while the animal tissue samples in general show a predominance and rise in 

BDE 47. The rise in PBDE concentrations in human blood and breast milk in North Americans 

(both from the United States and Canada) throughout the 1990s into the 2000s, coupled with the 

fact that North American body burdens exceed those of Europeans and others by factors of 10 or 

more, has served to focus attention on North American exposures to PBDEs. The pentaBDE 

formulation has garnered the most concern because it appears to be the major contributor to 

current environmental, biotal, and human body levels. Even with both the penta and octa 

formulations having been withdrawn from the U.S. market, past use and possible debromination 

of higher-brominated congeners (e.g., BDE 209) by photolytic or biological mechanisms to form 
xxiv

lower-brominated congeners might result in the continued presence of lower-brominated 

congeners in humans and the environment. 

Studies measuring the concentrations of the BDE congeners in environmental and 

exposure media concentrations were compiled and summarized with the ultimate goal of 

selecting representative BDE congener concentrations in exposure media (air, dust, food, etc.) to 

which Americans are exposed. While the data were insufficient to derive concentrations that 

could be considered statistically representative of the general U.S. population, they were deemed 

adequate for conducting the exposure modeling done in this report. These exposure media 

concentrations were combined with exposure contact rates (air breathed, food eaten, etc.) in 

order to estimate an overall intake dose. Concentrations and contact rates were used to 

characterize background central tendency exposures. Exposure media concentrations were either 

the straight average or geometric mean of the concentrations found in the study selected to 

represent national background conditions. Contact rates were arithmetic averages for adult 

populations as provided in EPA’s Exposure Factors Handbook. 

As lipophilic contaminants, PBDEs bioaccumulate in the lipids of organisms. This 

generalization is less true for BDE 209; it is not as lipophilic as other BDEs despite it being 

found and quantified in lipids of animals and humans. While the term “bioaccumulation” may 

also be questionable for BDE 209, there is little doubt that it occurs in human and animal tissues

(see discussion above on concentrations in animal tissues). In humans, lipid concentrations are 

typically measured in blood and breast milk, although adipose tissue concentrations have also 

been measured. Studies measuring BDE 209 in both blood and mother’s milk show lower 

concentrations in mother’s milk, suggesting limited transport for this congener within the body. 

Measured human concentrations of BDE congeners were compiled, and representative 

background concentrations of BDE congeners in human blood and milk of Americans were 

selected. Then, a simple pharmacokinetic model was used to predict the background lipid 

concentrations of the BDE congeners in humans using the background intake estimates. These 

predictions were compared to the selected representative body burden concentrations. The 

model predictions matched the measurements fairly well (as discussed below), suggesting 

confidence in the exposure characterization. Figure E-2 displays this study approach. 

 This exposure exercise focused on the congeners most often studied and found in the 

environment. It is noted, however, that there is not a final selection of toxic or otherwise most 
xxv

critical, congeners as there is, for example, with dioxin-like compounds. Some studies focus on 

as few as 4 congeners while others measure over 15 congeners. The total concentrations found 

in the individual studies (including for discussions below) can mean the sum of different

congeners. Many studies have focused on the penta formulation congeners and not measured the 

critical deca congener, BDE 209. Reasons cited for not including this congener include the 

historical predominance of the other congeners and the analytical difficulties associated with 

measuring BDE 209. The congeners selected for final modeling in this exercise include BDE 28, 

BDE 47, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, and BDE 209. 

Tables E-1 and E-2 provide the results of this exposure exercise. 

Figure E-2. Approach for characterizing exposure to polybrominated 

diphenyl ethers in this report. 
xxvi

Table E-1. Representative exposure media concentrations* 

Exposure Media Congener Number

28 47 99 100 138 153 154 183 209 Total

Water, pg/L 3.3 42.7 27.6 7.2 0.3 3.9 2.9 4.4 42.3 146.1

Surface soil, ng/g dwt --- 1.9 3.6 0.4 --- 5.7 4.8 37.4 15.3 82.3

Indoor dust, ng/g ND 1,857 2,352 911 181 243 156 60 2,394 8,275

Outdoor air, pg/m3 3 53 51 13 --- 4 4 1 25 158

Indoor air, pg/m3 27 177 79 16 --- 5 7 --- 121 447

Shellfish, ng/g wwt ND 3.6 1.2 0.9 ND ND ND ND ND 5.7

Finfish, ng/g wwt 0.01 0.20 0.04 0.04 0.0007 0.006 0.015 0.001 0.008 0.32

Beef, ng/g wwt 0.0006 0.009 0.01 0.001 ND 0.003 0.0009 0.0004 0.003 0.028

Pork, ng/g wwt 0.0007 0.031 0.033 0.005 0.0009 0.007 0.003 0.02 0.016 0.13

Poultry, ng/g wwt 0.0003 0.005 0.007 0.001 ND 0.001 0.001 0.001 0.004 0.020

Dairy, ng/g wwt 0.001 0.04 0.032 0.005 ND 0.004 0.002 0.001 0.675 0.79

Eggs, ng/g wwt ND 0.009 0.018 0.007 ND 0.007 0.004 ND 0.034 0.084

*Totals may include congeners other than those listed and are greater than the sum of the studied congeners; see 

Chapter 4 for more detail. 

--- Means no data available; ND = data available but not detected; and wwt = wet weight.
xxvii

Table E-2. Predicted adult doses and lipid-based concentrations of BDEs, 

with predicted concentrations compared with lipid-based measurements in 

blood and milk

Exposure Doses 

Pharmacokinetic 

Results 

Congener Number TOTAL 

28 47 99 100 138 153 154 183 209 

I. Exposure Doses

Soil ingestion, ng/d 0.00 81.25 102.92 39.86 7.92 10.67 6.86 2.86 104.83 357.17 

Soil dermal 

contact, ng/d 0.00 19.50 24.70 9.57 1.90 2.56 1.65 0.69 25.16 85.73 

Inhalation, ng/d 0.32 2.15 1.00 0.21 0.00 0.06 0.09 0.00 1.45 5.28 

Food & water 

ingestion, ng/d 0.22 18.98 7.98 4.39 0.03 0.70 0.46 0.66 10.75 44.17 

II. Pharmacokinetic Results

Predicted conc., 

ng/g lwt 0.1 8.6 12.5 3.9 1.3 3.9 0.6 <0.1 <0.1 31.0 

Observed blood, 

ng/g lwt 1.2 20.5 5.0 3.9 NA 5.7 NA NA NA 36.3 

Observed milk, 

ng/g lwt 1.7 26.0 5.4 5.2 NA 4.8 0.4 0.2 0.4 44.1 
xxviii

Table E-1 displays the final selection of congener concentrations in exposure media. 

Very few studies were found that characterized surface water and surface soil. Surface water 

concentrations, used as a surrogate for drinking water, were measured in the San Francisco 

estuary. The total water concentration was 146 pg/L. A total of 33 surface soil measurements 

were taken in 15 states and measured for 30 BDE congeners in the single study found for the 

United States. Concentrations of total BDEs averaged 82 ng/g dry weight (dwt), with a 

geometric mean concentration of 5.3 ng/g dwt. Average soil concentrations were used in this 

exercise. Outdoor air was characterized by a study by the California Air Resources Board 

(CARB). The CARB data of 84 samples were taken in 2004 from 7 monitors on 12 dates from 

locations in the Bay Area and the South Coast. While the profile with a 158 pg/m3 average 

might be higher than other outdoor profiles, it is still a reasonable representation of urban or 

suburban conditions. The indoor environment has been a focus of study because of the use of 

PBDEs in consumer products found in the home; a number of studies measuring PBDEs in 

indoor air and house dust were found. The profile in indoor house dust originated from a study 

from dust taken in 10 homes in 9 different states in the western part of the country. The total of 

8,275 ng/g dwt in that representative profile compares with 5,811 ng/g dwt from a study of 

17 homes in Washington, DC area, and 9,271 ng/g dwt from 2 samples from a computer lab in 

California. The geometric means of three locations (bedroom, living room, and from a 

household vacuum) within 20 homes in Boston of 6,332 ng/g, 13,882 ng/g, and 4,213 ng/g, 

respectively also compare well to the 8,275 ng/g dwt estimate. Based upon all these studies, a 

range of average concentrations in dust may be about 4000 to 14,000 ng/g dwt. BDE 209 

dominates the dust profile, and recent studies have found unusually high levels of BDE 209, 

including concentrations in the hundreds of thousands to over two million of ng/g dwt in the 

United Kingdom, and in the hundreds of thousands in one study in the United States. The indoor 

air profile came from this study of Boston homes in which the authors determined geometric 

mean concentrations for three locations (personal, bedroom, living room). The average of the 

three geometric means, 447 pg total BDEs/m3

, was used as the indoor air concentration. 

Numerous studies on concentrations in food were summarized. A wealth of studies on 

fish in the wild, raised in aquaculture, and purchased in the market place showed a wide range in 

concentrations. Generally, concentrations in the wild were found to be higher, and, in some 

studies, substantially higher than farm-raised or store-bought fish. Essentially all of the 
xxix

representative food profiles, including the fish profiles, originated from a single study sampling 

food from the retail market place in Texas. While limited geographically, the market basket 

survey includes measurement of 31 samples, each a composite of 10 samples of the same food 

type taken from markets in Dallas, Texas for a total of 310 samples collected, Food types 

sampled include meat, dairy, fish, and eggs. Thirteen congeners, including BDE 209, were 

measured. The average total concentrations (on a wet weight, not lipid weight, basis) used from 

this study include the following: 0.32 ng/g in finfish, 0.028 ng/g in beef, 0.13 ng/g in pork, 

0.020 ng/g in poultry, 0.79 ng/g in dairy, and 0.084 ng/g in eggs. A total concentration of 

5.7 ng/g in shellfish originated from the study in the San Francisco estuary, which also measured 

surface water concentrations. 

Body-burden studies were compiled and reviewed. Nearly all body-burden data are from 

human blood and breast milk. Table E-2 shows the final selected profiles of BDE congeners in 

blood and human breast milk of Americans. Body burdens of Americans are higher than body 

burdens of individuals in other countries; these data suggest total PBDE body burdens in the 

range of 30 to 100 ng/g lipid weight (lwt) for Americans, while body burdens of less than 

10 ng/g lwt have been found for people in other countries. Most of the data outside of the United 

States are from Europe. The study selected as the representative blood profile came from the 

National Health and Nutritional Examination Study (NHANES) from 2003/2004. The geometric 

mean total PBDE concentration of all adults from this study was 36 ng/g lwt; the 95th percentile 

was 291 ng/g lwt. BDE 209 was not quantified in individual samples of NHANES; hence, 

Table E-2 lists it as NA. However, in the literature publication on this data set, it was stated that 

BDE 209 concentrations in pooled samples from this set of NHANES samples had a 

concentration of 2 ng/g lwt. The predominant congener found in body burden studies is BDE 47, 

explaining about 50% of the total concentration. The next most abundant congeners are BDE 99 

and BDE 153, both explaining the range of 10−20% of total concentrations. Most of the studies 

have not measured BDE 209, including the NHANES study, but, when measured, it was found in 

about half the samples at low levels near 5 ng/g lwt, with the exception of one case. In this case 

study, which entailed taking measurements from a family of four, including two parents and two 

young children, concentrations of BDE 209 were above 100 ng/g lwt in the children. Low levels 

of BDE 209 have been attributed to the rapid half-life of 15 days in humans, and the higher 

levels in children from this study were attributed to dust exposures in the house. 
xxx

Unlike human blood data, there was no nationally representative breast milk study that 

could be used to characterize concentrations of PBDEs. Of the several studies evaluated, 

perhaps three studies could be used to represent background conditions. These include one by 

the Environmental Working Group, which sampled 20 women from around the United States; 

the Northwest Environment Watch (NEW) study, which had a sample size of 40 including 

women residing in several states in the Northwest; and perhaps a third study of 47 samples taken 

in Texas. The NEW study, which included measurements of BDE 209, was used. The median

total BDE concentration of the 40 samples was 44.1 ng/g lwt.

The first-order, single-compartment pharmacokinetic model used in this exercise to 

convert intakes to body lipid concentrations requires the half-life elimination rates of the BDE 

congeners and BDE absorption fractions (i.e., fractions of BDE intakes from ingestion of dust, 

water, and foods, as well as inhaled BDEs, that are absorbed in the stomach or the lungs to 

accumulate in body lipids) as inputs. Only two studies were found that provided human 

half-lives, with one deriving half-lives ranging from 2.9 to 11.0 years for congeners BDE 47, 

BDE 99, BDE 100, BDE 138, BDE 153, and BDE 154. The second reference derived 

significantly smaller half-lives for the higher-brominated congeners BDE 183 and 

BDE 209―half-lives of 0.26 year and 0.041 year, respectively. The absorption fraction for 

BDEs in dust ranged from 0.04 (BDE 209) to 0.78 (BDE 100). These were derived from a study 

of the bioavailability of BDEs in dust when fed to rats. The absorption fractions for food and 

water were near 0.90 for all of the congeners. The soil-dermal contact pathway required a 

fraction of PBDE absorbed through the skin, set at 0.03 for all congeners, as well as parameters 

reflecting adherence of soil contacted by the skin, contact events, and contact surface areas. 

Table E-2 provides the final results from this exposure assessment. The exposure 

pathways of dust ingestion and dust-dermal contact dominated total adult exposure, explaining 

90% of the 492 ng/day total (equivalent to 7 ng/kg-day for the 70 kg adult). The congeners 

BDE 47, BDE 99, and BDE 209 dominated the exposure intake: each explained between 25 and 

29% of total intakes, with other congeners explaining the remaining intakes. Using the estimated 

adult dose of 492 ng/day, a blood serum concentration of BDE was predicated. The predicted 

concentration of total BDEs of 31.0 ng/g lwt was similar to the NHANES human blood 

measurement of 36.4 ng/g lwt and the breast milk concentration of 44.1 ng/g lwt. Predictions 

appear reasonably close to measurements for six of nine congeners. However, predicted 
xxxi

concentrations of BDE 47 and 209 were less than NHANES measurements and the predicted 

concentration of BDE 99 was more than NHANES measurements. The cause of the differences 

between predicted and measured values for these congeners was unknown, but could point to the 

inappropriate assignment of elimination half-life, or in the case of BDE 99, metabolic 

debromination. 

While predictions of adult body burdens were close to observations, uncertainties exist in 

the intake dose estimates and the pharmacokinetic modeling, starting from development of dose 

estimates based on limited environmental measurements, to indoor contact rates with house dust, 

to the pharmacokinetic parameters of absorption and elimination half-life. Contact rates for 

food/water ingestion and inhalation are fairly well established, and the exposure media 

concentration summaries suggest similarities among different studies. Consequently, the dose 

via food/water ingestion and inhalation might be considered reasonably certain, for purposes of 

this discussion. However, food/water ingestion and inhalation explained less than 20% of the 

body burden, based upon the estimate of total exposure derived using the pharmacokinetic 

model. The remainder of the estimated exposure likely came from house dust through the 

pathways of ingestion and dermal contact, or some other, unknown source. Assignment of dust 

contact rates in combination with concentrations found in house dust lead to exposures that 

resulted in reasonably accurate predictions of measured body burdens. The assertion that dust is 

a major source of PBDE exposure is supported by the circumstantial evidence that PBDE 

concentrations in U.S. house dust are generally ten times higher than concentrations reported in 

European studies and that European body burdens (ca 10 ng/g lwt) are as much as ten times 

lower than U.S. body burdens (30 – 100 ng/g lwt). 

In addition to adult body burden measurements and intake dose estimates, exposures for 

other populations were also examined. These unique exposures included fetal exposures, infant 

exposures via breast milk, childhood exposures, occupational, and unusually high exposures at 

the high end of the general population. Key findings from these examinations are summarized 

below. 

PBDE Body Burdens in U.S. Children: Body burden data, as well as intake and body 

burden impact modeling, suggest that infants and toddlers have higher exposures than older 

children or adults. In paired sampling studies, including mother and child, the child’s body 

burden has exceeded by the mother’s by about a factor of 4, exceeding 100 ng/g lwt total PBDEs 
xxxii

in some cases. Higher levels in toddlers were attributed to ingestion of mother’s milk and high 

exposures to house dust. 

PBDE Fetal Body Burdens: Limited data from fetal tissue and numerous studies 

including measurements of BDE congeners from umbilical cord blood support the conclusion 

that the fetus is exposed to PBDEs through the mother. In some paired studies including both 

maternal and umbilical cord serum, there is a suggestion that the fetus may be less exposed than 

the mother as lower concentrations of BDE congeners in umbilical cord as compared to maternal 

serum were found. This trend of lower concentrations were seen mostly for higher-brominated 

congeners in two of these studies, where this general observation could be made, but in four 

other studies reviewed, similar, if not slightly higher, concentrations of all congeners were seen 

in umbilical cord serum.

Intake Rates for Infants and Children: Intake modeling for the breast milk pathway 

combining measured milk concentrations and infant ingestion of human milk led to an intake of 

1,411 ng/day in this study. Assuming an average body weight of 10 kg for an infant during the 

months of breast-feeding, a dose is calculated as 141 ng/kg-day. It is lower than intake estimates 

for children derived in this study, which were 47.2 ng/kg-day for ages 1−5 (assuming 15 kg bw), 

13.0 ng/kg-day for ages 6−11 (assuming 30 kg bw), and 8.3 ng/kg-day for ages 12−19 (assuming 

58 kg bw). The estimated adult intake rate was 7.1 ng/kg-day. 

Occupational Exposure: Limited occupational data support the observation that 

individuals in occupations that would lead to higher exposures to specific congeners have higher 

concentrations of PBDE congeners in their blood than the general population. 

High-End Exposures in U.S. Population: Even in background adult populations, there 

are individuals experiencing very high exposures. As noted earlier, a range of body burdens of 

total PBDEs found in the literature was 30-100 ng/g lwt, and the 95th percentile found in 

NHANES was 291 ng/g lwt. Modeling exercises undertaken to examine this hypothesis suggest 

that even the highest dust concentrations might not be able to explain the highest body burdens 

found. 

This report has quantified central tendency intake estimates and body burdens of PBDEs, 

and found that dust-related exposures best explain body burdens. However, this examination of 

unusually high exposures in the general population suggests the possibility that there are other 

exposures not identified in this assessment, or exposures that were identified but not quantified 
xxxiii

properly for all individuals of the population. These exposures not identified or improperly 

quantified are affecting a small percentage of individuals at the high end (upper 5%, perhaps) of 

the U.S. population. This might be the most important uncertainty identified in this report. 
1-1

1. INTRODUCTION

 Polybrominated diphenyl ethers (PBDEs) are a class of brominated flame retardants that 

are added to plastics, polyurethane foam, textiles, and electronic equipment to reduce the 

likelihood of ignition and to slow the burn rate if the products do catch fire. PBDEs have a 

common structure of a brominated diphenyl ether molecule that may have anywhere from 1 to 

10 bromine (Br) atoms attached. Depending on the location and number of Br atoms, there are 

209 possible PBDE compounds, each termed congeners. Each is assigned a specific brominated 

diphenyl ether (BDE) number (note: in this document, the abbreviation PBDE was used to 

denote the class of brominated flame retardants, while BDE was used in the context of PBDE 

congeners). For example, there are 42 tetrabromodiphenyl ether congeners (those with four 

bromine atoms), but only a few of them, specifically BDE 47 and occasionally BDE 66, are 

found in the product formulations and in environmental or exposure media (La Guardia et al., 

2006). Table 1-1 shows the BDE congener number and chemical composition of the most 

commonly studied BDE congeners. 

 PBDEs have been marketed in three primary formulations: the pentaBDE formulation, 

commercially known as DE-71 and Bromkal 70-5DE; the octaBDE formulation, DE-79; and the 

decaBDE formulation, DE-83R or Saytex 102E. The formulations differ in their composition of 

BDE congeners. The penta formulation is dominated (by weight) by penta congeners (50−62%) 

with secondary contributions by tetra (24−38%) and hexa congeners (4−12%). The octa 

formulation is dominated by hepta (45%) and octa congeners (33%), with secondary 

contributions by hexa (12%) and nona (10%) congeners. The deca formulation is composed of 

essentially all BDE 209 (97−99%, with 1−3% other, mainly nona, congeners), which is the 

congener with all 10 Br positions occupied. The penta and octa formulations were voluntarily 

withdrawn from the U.S. marketplace by their manufacturers at the end of 2004, leaving only the 

deca formulation for use in commercial products. The deca formula is also the only currently 

used formulation in Europe because the penta and octa formulations were banned. However, 

Sweden banned the use of the deca formulation in August of 2006, the ban to take effect 

1/1/2007.1

 

1

See http://www.emfacts.com/weblog/index.php?p=547. 
1-2

Table 1-1. BDE congener numbers and chemical composition of the most 

commonly studied BDE congeners

BDE congener number Chemical formula BDE congener number Chemical formula

I. MonoBDE BDE 118 2,3',4,4',5-BDE

BDE 3 4-BDE BDE 119 2,3’,4,4’6-BDE

II. DiBDE BDE 126 3,3’,4,4’,5-BDE

BDE 7 2,4-BDE BDE 138 2,2’,3,4,4’,5’-BDE

BDE 8 2,4’-BDE BDE 140 2,2’,3,4,4’,6-BDE

BDE 11 3,3’-BDE VI. HexaBDE

BDE 12 2,6-BDE BDE 153 2,2’,4,4’,5,5’-BDE

BDE 13 3,4’-BDE BDE 154 2,2’,4,4’,5,6’-BDE

BDE 15 4,4’-BDE BDE 155 2,2’,4,4’,6,6’-BDE

III. TriBDE BDE 166 2,3,4,4',5,6-BDE

BDE 17 2,2’,4-BDE VII. HeptaBDE

BDE 25 2,3’,4-BDE BDE 181 2,2’,3,4,4’,5,6-BDE

BDE 28 2,4,4’-BDE BDE 183 2,2’,3,4,4’,5’,6-BDE

BDE 30 2,4,6-BDE BDE 190 2,3,3’,4,4’,5,6-BDE

BDE 32 2,4’,6-BDE VIII. OctaBDE

BDE 33 2’,3,4-BDE BDE 196 2,2’,3,3’,4,4’,5’,6-BDE

BDE 35 3,3’,4-BDE BDE 197 2,2’,3,3’,4,4’,6,6’-BDE

BDE 37 3,4,4’-BDE BDE 203 2,2’,3,4,4’,5,5’,6-BDE

IV. TetraBDE IX. NonaBDE

BDE 47 2,2’,4,4’-BDE BDE 206 2,2’,3,3’,4,4’,5,5’,6-BDE

BDE 49 2,2’,4,5’-BDE BDE 207 2,2’,3,3’,4,4’,5,6,6’-BDE

BDE 66 2,3’,4,4’-BDE BDE 208 2,2’,3,3’,4,5,5’,6,6’-BDE

BDE 71 2,3’,4’,6-BDE X. DecaBDE

BDE 75 2,4,4’,6-BDE BDE 209 2,2’,3,3’,4,4’,5,5’,6,6’-BDE

BDE 77 3,3’,4,4’-BDE

V. PentaBDE

BDE 85 2,2’,3,4,4’-BDE

BDE 99 2,2’,4,4’,5-BDE

BDE 100 2,2’,4,4’,6-BDE

BDE 105 2,3,3’,4,4’-BDE

BDE 116 2,3,4,5,6-BDE
1-3

 PBDEs have captured the attention of scientists and policymakers because levels in the 

environment and humans have increased rapidly since these chemicals came into use. The rise in 

PBDE concentrations in blood and breast milk samples both from the United States and Canada 

throughout the 1990s into the 2000s, coupled with the fact that North American body burdens 

exceed those of Europeans and others by factors of 10 or more, has served to focus attention on 

North American exposures to PBDEs. The pentaBDE formulation has garnered the most 

concern because it appears to be the major contributor to current environmental and human body 

levels. Even with both the penta and octa formulations having been withdrawn from the 

U.S. market, past use and the possibility of debromination (loss of bromine atoms) of BDE 209 

and other higher-brominated congeners by photolytic or biological mechanisms to form 

lower-brominated congeners could result in the continued presence of lower-brominated 

congeners in the environment. 

 Studies have been conducted in laboratory animals to gain a better understanding of the 

potential health risks of PBDEs. These studies have suggested potential concerns about liver 

toxicity, thyroid toxicity, developmental toxicity, and developmental neurotoxicity―especially 

about potential risks in children. The carcinogenic potential of some PBDEs have been studied. 

In a review of toxicological studies as part of a toxicological review for EPA’s Integrated Risk 

Information System (U.S. EPA, 2006), EPA has found that the data for decabromodiphenyl ether 

(the commercial formulation composed mostly of BDE 209) support a finding of “suggestive 

evidence of carcinogenic potential” according to EPA’s Guidelines for Carcinogen Risk 

Assessment (U.S. EPA, 2005). For BDE congeners BDE 99, BDE 153, and BDE 47, EPA has 

found that the data support a finding of “inadequate information to assess carcinogenic potential” 

(U.S. EPA, 2005). 

 EPA has formed a working group composed of individuals from several program offices 

including the Offices of Pesticides, Prevention, and Toxic Substances, the Office of Water, and 

of the Office of Research and Development, to study issues surrounding PBDEs.

2

 

2

See http://www.epa.gov/oppt/pbde. 

 They issued a 

project plan in 2006 that outlined projects to further study the toxicity, environmental fate, and 

exposure to PBDE compounds. This document addresses the exposure assessment needs 

identified in that project plan and provides a comprehensive assessment of the exposure of 
1-4

Americans to this class of persistent organic pollutants. That document was updated in 

December of 2008 and a further update is planned for 2010. 

 Subsequent chapters describe the historical use and composition of commercial mixtures 

of PBDEs (see Chapter 2), the environmental fate of PBDEs (see Chapter 3), and environmental 

and exposure media concentrations of PBDEs (see Chapter 4). The document concludes with a 

comprehensive exposure assessment addressing fetal, infant, children, and adult exposures to 

PBDEs (see Chapter 5). 

REFERENCES FOR CHAPTER 1

La Guardia, MJ; Hale, RC; Harvey, E. (2006) Detailed polybrominated diphenyl ether (PBDE) congener 

composition of the widely used penta-, octa-, and deca-PBDE technical flame-retardant mixtures. Environ Sci 

Technol 40:6247−6254.

U.S. EPA (Environmental Protection Agency). (2005) Guidelines for carcinogen risk assessment. Fed Reg 

70(66):17765−18717. Available online at http://www.epa.gov/cancerguidelines. 

U.S. EPA (Environmental Protection Agency). (2006) Toxicological review of decabromodiphenyl ether (BDE 209) 

(CASRN 1163-19-5) in support of summary information on the Integrated Risk Information System (IRIS). 

December, 2006. Available online at http://www.epa.gov/iris. 
2-1

2. PRODUCTION, USE, AND LIFE CYCLE OF POLYBROMINATED

DIPHENYL ETHERS

2.1. INTRODUCTION

Polybrominated diphenyl ethers (PBDEs) are in widespread use and are now detected in 

the tissues of humans and wildlife, in soils, in sediments, and in air. The ubiquitous presence of 

brominated flame retardants in the environment has heightened concerns for adverse ecological 

and human health risks. PBDEs are a class of aromatic compounds intentionally manufactured 

to retard the combustibility of treated materials. When fire occurs, the PBDE formulations 

utilize vapor phase chemical reactions that interfere with the combustion process, thus delaying 

ignition and inhibiting the spread of fire (D’Silva, 2004). These characteristics have promoted 

the widespread use of PBDEs in textiles, flexible polyurethane foams (FPUFs) used in 

upholstery stuffing for furniture and car seats, electronic components, electrical components, and 

plastics used in the casings of televisions (TVs), personal computers (PCs), and other electronic 

equipment. 

The purpose of this chapter is to describe the production and uses of PBDEs, and to 

present a basic life cycle analysis of PBDEs involving estimates of environmental release from

production, product use, waste disposal, recycling of electronics and electrical equipment (EEE) 

materials, and sewage treatment. Section 2.2 reviews the production of commercial PBDE 

formulations. Section 2.3 surveys the specific uses of PBDEs as flame retardants in a number of 

plastic resins and finished products. Section 2.4 presents a generalized life cycle analysis for the 

United States that derives estimates of environmental releases of specific brominated diphenyl 

ether (BDE) congeners based on production, the uses of PBDE-treated products, and estimates 

the amounts that may be recycled or disposed of in landfills and incinerators once PBDE-treated 

products become functionally obsolete. Section 2.5 presents a summary and the key findings 

from this life cycle analysis. Because of significant limitations and gaps in existing data, the life

cycle analysis presented in this chapter should be regarded as an approximate assessment. 

However, the life cycle analysis does provide a platform from which distinct observations can be 

made regarding the potential pathways of environmental releases of PBDE formulations and 

their BDE congeners into the land, air, and water of the United States from production through 

disposal of PBDE-treated materials.
2-2

2.2. PRODUCTION

Commercial production of PBDEs began in 1976 (IPCS, 1994). PBDEs have been sold 

under various trade names but mainly consist of three commercial mixture formulations known 

as penta-, octa-, and decaBDE. Each commercial formulation is manufactured through the 

chemical reaction of bromine with diphenyl oxide, (also known as diphenyl ether) in the 

presence of an inorganic catalyst (e.g., AlCl3) (ATSDR, 2004). The bromine amount and the 

time allotted for the chemical reaction control the extent of bromination on the diphenyl ether 

molecule. The stepwise addition of bromine causes the formation of lower to higher-brominated

PBDE congeners until the total desired amount of bromination is obtained. Figure 2-1 displays 

the general structure of the PBDE compound. The molecular backbone consists of two phenyl 

rings interconnected by an oxygen atom. There are 10 positions whereby a bromine atom can 

substitute a hydrogen atom on the molecule with the possibility of 10 homologue groups 

identified by the preface mono-, di-, tri-, tetra-, penta-, hexa-, hepta-, octa-, nona-, and decaDBE. 

Each congener is indicated by the positional numbering of the bromine atom on the biphenyl 

rings. The congeners are assigned a number according to the nomenclature of the International 

Union of Pure and Applied Chemistry. 

Figure 2-1. General structure of polybrominated diphenyl ethers.

Commercial formulations are mixtures of PBDE congeners with penta-, octa-, and 

decaBDE having a bromine content of about 70.8%, 79%, and 83%, respectively (EU, 2001, 

2002, 2003). Although 209 PBDE congeners are theoretically possible, only a limited number 

have been detected in commercial flame retardant products. Table 2-1 is a compilation of the 
2-3

Table 2-1. Approximate BDE homologue- and congener-specific content of 

PBDE commercial formulations

Congener identity 

Commercial formulations of PBDE fame retardants

PentaBDE 

(mass % composition)

OctaBDE

(mass % composition)

DecaBDE

(mass % composition)

BDE 17 <0.1

BDE 28 0.2

Total TriBDE 0−1 0.00001

BDE 47 25−37 0.00003

BDE 66 <1

Total TetraBDE 24−38 0.00003

BDE 85 2

BDE 99 35−50 0.002

BDE 100 6−10

Total PentaBDE 50−62 0.002

BDE 138 0.5

BDE 153 3−5 5−10 0.001

BDE 154 2−4 1−5

Total HexaBDE 4−12 10−12 0.001

BDE 183 40

BDE 190 1

Total HeptaBDE 43−58 0.003

BDE 196 8

BDE 197 21

BDE 203 5−35

Total OctaBDE 26−35

BDE 206 2.2

BDE 207 7 0.24

BDE 208 10 0.06

Total NonaBDE 8−14 −2.5

BDE 209 0−3 97.5

Sources: EU (2001, 2002, 2003); Kemmlein et al. (2005); Peele (2004); Palm et al. (2004); IPCS (1994); 

La Guardia et al. (2006).
2-4

approximate BDE congener compositions of the three commercial formulations. The 

composition of the three commercial formulations change with time and manufacturer, and 

Table 2-1 shows general congener composition. The dominant congeners in pentaBDE (percent 

weight basis in parenthesis) are BDE 99 (35−50%), BDE 47 (25−37%), BDE 100 (6−10%), 

BDE 153 (3−5%), and BDE 154 (2−4%). On a homologue basis, the pentaBDE formulation is 

dominated by pentaBDE (50−62%) and tetraBDE (24−38%) homologues, with hexaBDE 

amounting to only 4−12% of total mass. 

The BDE congeners present in octaBDE formulations are BDE 183 (40%), BDE 197 

(21%), BDE 203 (5−35%), BDE 196 (8%), BDE 208 (10%), BDE 207 (7%), BDE 153 (5−10%), 

BDE 154 (1−5%), BDE 209 (not detected up to 3%), and BDE 190 (1%). Hexa-, hepta-, octa-, 

and nonaBDE homologues dominate the octaBDE commercial formulations comprising 

10−12%, 43−58%, 26−35%, and 8−14% of total BDE content, respectively. With respect to 

decaBDE, 97.5% of the formulation is composed of BDE 209, with the remainder being 

BDE 206 (2.2%), BDE 207 (0.24%), and BDE 208 (0.06%). BDE 209 is the dominant 

congener, and is a signature of this formulation. The congener composition of the commercial 

decaBDE product has become progressively dominated by the BDE 209 congener over time. 

For example, in a 1986 toxicity evaluation, the National Toxicology Program noted that 

BDE 209 was approximately 96% (or greater) in the commercial decaBDE product (NTP, 1986). 

Because of mergers and acquisitions, PBDE flame retardant formulations are produced 

worldwide by three chemical manufacturers: Albemarle, Chemtura, and ICL Industrial Products. 

In the United States, all PBDE flame retardant products sold are currently decaBDE

formulations. Prior to 2005, the industry was more diverse. The Dead Sea Bromine Group 

(DSBG), based in Israel, was a leading manufacturer of PBDE-flame retardants. ICL Industrial 

Products acquired the DSBG in 2005. The major producer of PBDE products in the United

States was the Great Lakes Chemical Corporation, which sold flame retardants under the 

Firemaster brand name. The Great Lakes Chemical Corporation merged with Crompton 

Chemical to form Chemtura. Chemtura produces flame retardant products composed of 

decaBDE under the brand names AZUB DB-40, AZUB DB-65, AZUB 2DA-65, and 

AZUB 3DA-65. Prior to 2005, the Great Lakes Chemical Corporation produced a 

pentaBDE-based product under the brand name DE-60F. DE-60F was replaced with 

Firemaster-550, a phosphorous-bromine-based flame retardant. Firemaster-550 is technically not 
2-5

enriched with PBDE, but the exact active chemical ingredients remain proprietary. However, 

Stapleton et al. (2008) analyzed dust samples in homes in the Boston area and identified these 

brominated components of Firemaster-550: 2-ethylhexyl 2,3,4,5-tetrabromobenzoate (TBB) and 

(2-ethylhexyl)tetrabromophthalate (TBPH), the latter compound being a brominated analogue 

of di(2-ethylhexyl)phthalate (DEHP). Albemarle produces a decaBDE-formulated flame

retardant under the brand name SAYTEX 102E Flame Retardant. Because of increased 

international regulatory concern regarding the potential for human health and environmental 

effects, the PBDE industry voluntary ceased production of penta- and octaBDE on 12/31/2004 

(Birnbaum and Hubal, 2006). The EPA announced the phase-out of decaBDE in December of 

2009 (U.S. EPA, 2009c). The three companies producing PBDEs (noted above) have committed 

to end production, importation, and sales of decaBDE for most uses in the United States by 

12/31/2012, and to end all uses by the end of 2013. 

Production figures for PBDE manufacture are not readily available. The PBDE industry 

usually reports only on market demand for PBDE commercial formulations, which likely 

translates into sales volume. The latest reporting year for market demand is 2001 (Hites, 2004). 

The market demand for decaBDE in 2001 within the Americas was approximately 24,500 metric 

tons. The Americas includes North, Central, and South America, combined. There are no 

further breakdowns available by country. In 2001, 83% of all PBDE consumed worldwide was 

decaBDE, followed by pentaBDE (11%), and octaBDE (6%). Within the Americas, the 

breakdown of the market demand for deca-, penta-, and octaBDE in 2001 was 74%, 21.5%, and 

4.5%, respectively (analysis of data in Hites, 2004). Two-thirds of all decaBDE produced 

globally are consumed in North America and Asia (analysis of data in Hites, 2004). 

Concerns for environmental persistence, the potential for bioaccumulation into ecological 

and human food chains, and the widespread occurrence of PBDEs in the environment have 

precipitated regulatory and voluntary actions to further reduce the use of PBDEs. The sole 

U.S. manufacturer of penta- and octaBDE phased out production of these products at the end of 

2004. Additionally, several states enacted individual legislation banning the use of penta-, octaand decaBDE formulations as summarized below. 

• As of 6/1/2006, California and Maine prohibited the manufacture, distribution, and 

processing of products containing penta- and octaBDE flame retardants (CalEPA, 2006; 

NCEL, 2005). 
2-6

• As of 1/1/2006, Hawaii, Illinois, and Michigan prohibited the manufacture, processing, 

and distribution of products containing more than 0.1% of the PBDE formulations pentaand octaBDE (Illinois EPA, 2006; NCEL, 2005). 

• Minnesota similarly prohibited the manufacture and uses of penta- and octaBDE effective 

1/1/2008 (NCEL, 2007). 

• Washington State banned the uses of penta-, octa-, and of decaBDE in mattresses 

effective 1/1/2008 (Washington State, 2007). However, the Washington State regulation 

does not prohibit the use of decaBDE in residential upholstered furniture, televisions, or 

computers with electronic enclosures until 1/1/2011. 

• Beginning 10/1/2008, Maryland will prohibit the manufacture, processing, and 

distribution of products containing more than 0.1% of the PBDE formulations of pentaand octaBDE (NCEL, 2007). 

• As of August 2004, New York prohibited the manufacture and processing, but not the 

sale or use, of penta- and octaBDE (NCEL, 2005). 

Internationally, other countries and governmental entities have banned the use of penta-, 

octa-, and decaBDE. The European Union (EU) banned the marketing and use of penta- and 

octaBDE effective 2/6/2003, with Directive 2003/11/EC of the European Parliament and Counsel 

(European Parliament, 2003). At the time, the EU exempted decaBDE from the ban. However,

on 4/1/2008, the European Court of Justice (the EU’s highest court) ruled that this decision to 

exempt decaBDE was flawed and ordered the EU to ban the use of the decaBDE in electrical and 

electronic equipment (European Court of Justice, 2008). Based on the directive of the court, the 

ban became effective on 7/1/2008. The ban applies to all EU member countries. Australia

banned the use and import of pentaBDE effective 3/6/2007 (Australian Government, 2007). 

However, it is expected that the use of recycled plastics containing PBDE formulations may 

continue to incorporate these chemicals in the manufacture of new electronics and information 

technology (IT) units (Morf et al., 2005). 

2.3. USES OF PBDEs 

Due to the phase out by the manufacturer, as well as the state-specific bans noted above, 

penta- and octaBDE are no longer used in products to inhibit flammability. For the most part, 

decaBDE is not prohibited from production and continues to be used. This section reviews the 

past uses of penta- and octaBDE, and the current uses of decaBDE. 
2-7

Prior to the end of production in 2004, approximately 95% of pentaBDE was used as an 

additive flame retardant in FPUF materials (EU, 2001). The treated FPUF was used as seat 

cushioning and backing material for domestic furniture; in bedding mattresses, and in cushioning 

for automobile seats and laminated automobile seat headrests (ATSDR, 2004). The remaining 

5% use of pentaBDE was in the treatment of foam-based packaging materials and carpet 

padding. Only 7.5% of the approximate 953,000 metric tons (MT) of FPUF produced annually 

in the United States was treated with pentaBDE (ATSDR, 2004). Typically, the pentaBDE was 

mixed with aromatic phosphate esters in a ratio of 3:1 prior to application to FPUF. Other past 

uses of pentaBDE were in textile fabrics used in upholstery for furniture and automobile seat 

covers; epoxy resins (EPs) used as protective coatings on circuit boards; unsaturated polyesters; 

paper laminates; flexible polyvinyl chloride used as electrical wire coatings; rubber; paints and 

lacquers; rigid polyurethane foam; and adhesives. However, pentaBDE was not added to 

acrylonitrile-butadiene-styrene (ABS)-based plastics used in the manufacture of casings to 

television sets, computers, hairdryers, and automotive parts (IPCS, 1994; ATSDR, 2004; 

Rahman et al., 2001). 

 Mixed with antimony trioxide, octaBDE was primarily used as an additive flame 

retardant for certain plastics. Approximately 95% of the use of octaBDE was as an additive 

flame retardant in the production of ABS-based plastics, with the remaining 5% used as an 

additive to high impact polystyrene (HIPS), polybutylene terephthalate (PBT), polyamide 

polymers, polycarbonate, nylon, polyolefin, and phenol-formaldehyde resins (ATSDR, 2004; 

EU, 2003; IPCS, 1994). OctaBDE was typically added to ABS at a loading of 12−15% weight 

of the final product (EU, 2003). Materials that were treated with octaBDE included the housings 

of office equipment and business machines, and PC casings (EU, 2003).

DecaBDE is used as a general purpose additive flame retardant to a wide array of plastics 

having many product applications. HIPS, polyethylene (PE), polypropylene (PP), PBT, and 

unsaturated polyesters (UPEs) are common plastics treated with decaBDE (Alaee et al., 2003; 

BSEF, 2006). A major use of decaBDE in the United States is as an additive flame retardant to 

HIPS. HIPS-based plastics are used in the manufacture of housings and back panels to 

televisions, in casings of audio and video equipment, mobile phones, remote controls, PCs, and 

PC monitors. The PE-based plastics are used in the insulation of wire and cables of electrical 

equipment. The PP-based plastics are used in communication cables, capacitor films, building 
2-8

cables, pipes, stadium seats, lamp sockets and holders, and kitchen hoods. The PBT-based 

plastics are used as connectors in electrical and electronic equipment. UPE is used in building 

and construction materials as reinforced plastic panels. DecaBDE is also added as a flame 

retardant to nylon, and to the back coating of upholstery textiles used in sofas, chairs, and office 

furniture. The decaBDE-treated nylon is used for connectors in electrical and electronic 

equipment, circuit breakers, and coils. DecaBDE is added to polymers in concentrations of 

10−15%, by weight, and in conjunction with antimony trioxide (Directorate-General 

Environment, 2005). A significant use of decaBDE in the United States is for the flame

retardation of the back panels and the HIPS-based plastic casings of television sets. The Lowell 

Center for Sustainable Production at the University of Massachusetts estimated that 

approximately 17,150 MT of decaBDE was used in 28 million television sets sold in the United 

States in 2003 (Pure Strategies, Inc., 2005). 

2.3.1. PBDE Content of Consumer Products

An assessment of the penta- and octaBDE content of various polymers was conducted by 

the German Environmental Protection Agency (Kemmlein et al., 2003). A total of 32 samples, 

8 each of ABS copolymer ABS resin, HIPS, polyurethane hard foam (PUR-H), and epoxy resin 

were evaluated using single-ion monitoring and a gas chromatographic/mass spectrometer 

system. The polymers were obtained from the manufacturers along with the percent addition of 

penta- and octaBDE as flame retardants. ABS samples had been treated with 1% pentaBDE and 

2.95% octaBDE; HIPS samples with 2.96% octaBDE; and PUR-H and EP with 2% pentaBDE. 

Table 2-2 shows the mean and standard deviation of the concentration of penta- and octaBDE 

detected in eight samples of each polymer. 

 A Swiss study published in 2005 provides a quantitative basis for assigning a plausible 

PBDE content of final products made from flame retarded plastic resins (Morf et al., 2005). In 

this study, waste electronic products and components were sampled at an electronics recycling 

facility in Bern. The sampled end products included TVs, video camcorders, radios, high fidelity 

(Hi-Fi) stereo systems, portable compact disc players, mobile phones, standard telephones, 

toasters, and vacuum cleaners. Together, these product lines represented about 90% of the total 

use of penta-, octa-, and decaBDE employed as flame retardants in EEE (Morf et al., 2005). 
2-9

Table 2-2. Mean concentration (mg/g) of penta- and octaBDE in various 

flame retarded polymers

Polymer PentaBDE OctaBDE

ABS 1.187 (±0.058) 0.528 (±0.035)

HIPS 0 1.057 (±0.061)

PUR-H 1.414 (±0.053) 0

EP 1.414 (±0.148) 0

ABS a acrylonitrile-butadiene-styrene-copolymer; HIPS = high-impact 

polystyrene; PUR-H = polyurethane hard foam; EP = epoxy resin.

Source: Kemmlein et al. (2005). 

Several samples of discarded material were taken from piles of copper cable, printed circuit 

boards, TV housings and rear covers, and PC housings. The samples from each product were 

composited and then analyzed using either gas chromatography or electron capture detection 

with mass spectrometry. The mean concentrations (mg/kg) of the PBDEs found in each product 

are shown in Table 2-3. 

Table 2-3. Mean concentrations of PBDEs (mg/kg) found in EEE waste 

material at a recycling plant in Switzerland

Product PentaBDE OctaBDE DecaBDE

Copper cable 25 (±10) 100 (±150) 170 (±110)

Printed circuit boards 17 (±7) 10 (±1) 27 (±9)

TV housings (wood) 10 (±4) 10 (±4) 20 (±30)

TV/PC housings (plastic) 50 (±3) 7,500 (±600) 4,800 (±400)

TV housing rear covers 50 (±20) 7,700 (±3,600) 13,000 (±9,000)

Source: Morf et al. (2005).

 In addition to the mean concentrations in EEE waste, Morf et al. (2005) determined the 

abundance of select PBDE congeners in the EEE materials. One sample of each of the following 
2-10

items were taken from the output piles of sorted materials at the recycling plant: a plastic PC 

screen housing, a plastic TV-housing rear cover, a fine-grained plastic with particle size 2−5 and 

5−10 mm (general EEE waste), a fine-grained metal, a fine particulate, and a printed circuit 

board. Samples were Soxhlet-extracted and analyzed with high resolution gas chromatography 

(HRGC)/high resolution mass spectrometry (HRMS). Table 2-4 summarizes the results of the 

analysis of these single EEE waste material samples. Because these were essentially grab 

samples, the results in Table 2-4 only serve as a general indication of the potential congener 

distribution in the EEE waste. 

Table 2-4. BDE congener concentrations (mg/kg) present in EEE waste 

components sampled at a recycling facility in 2002 

BDE 

congener

PC screen 

housings

(mg/kg)

Television 

housings 

rear covers

(mg/kg)

Fine 

particulates 

(dust)

(mg/kg)

Fine-grained 

plastics

(5−10 mm)

(mg/kg)

Printed 

circuit 

boards

(mg/kg)

BDE 28 0.32 0.32 6.5 0.47 1.3

BDE 47 3.0 2.2 39 5.1 9.6

BDE 99 6.5 3.8 52 7.5 12

BDE 100 0.65 0.28 4.4 0.78 0.72

BDE 153 598 450 31 110 4.0

BDE 154 66 38 6.9 12 1.1

BDE 183 3,800 3,900 150 690 12

BDE 209 7,300 13,000 760 2,500 89

PentaBDE 13 7.9 120 17 28

OctaBDE 11,000 11,000 420 2,000 34

DecaBDE 7,300 13,000 760 2,500 89

Source: Morf et al. (2005). The concentrations of technical BDE mixtures were calculated by Morf et al. (2005) as 

follows: pentaBDE: Σ (BDE 47 + BDE 99 + BDE 100) × 1.26 octaBDE: BDE 183 × 2.87 

decaBDE = BDE 209.
2-11

 The Consumer Products Safety Commission reported on the chemical flame retardant 

content of polyurethane foams (PUFs) used in furniture in the mid-1990s (Cobb, 2005). Seven 

PUF types were evaluated. Three samples of each of the foam types were collected and analyzed 

by HRGC/HRMS. Only one PUF material contained PBDE (PBDE species not identified). 

Typically, PUF was made flame retardant with pentaBDE up until its discontinued use in 2004 

(EU, 2002). An average of 3% (range 2.9−3.2%) PBDE was found in this particular PUF. The 

other samples contained melamine and/or TDCP as flame retardants. Hale et al. (2002) noted in 

a separate study that the pentaBDE content in PUF can be as high as 30% by weight. 

2.4. GENERALIZED LIFE CYCLE OF PBDEs 

This section presents a general analysis of the life cycle of PBDE as it relates to 

production, use, recycling, and disposal in the United States. In general, the life cycle of PBDE, 

from production to eventual disposal, includes activities that can cause environmental releases of 

the chemical. It begins with the production of PBDE commercial formulations, then proceeds to 

applications to textiles and polymers, to polymer applications in consumer goods, to the use of 

consumer goods treated with PBDE flame retardants, and to the final disposal and recycling of 

PBDE-treated consumer goods as products reach obsolescence. Figure 2-2 portrays the idealized 

life cycle of PBDE. Each element of the life cycle can potentially release PBDE to the open and 

circulating environment. If all data were available to address each facet of the life cycle, then the 

life cycle would be considered complete. However, with regard to all opportunities for 

environmental releases, the scarcity of information restricts the life cycle analysis presented in 

this chapter, limiting what can be said about potential releases during production, the consumer 

use of PBDE-treated products, and the disposal and recycling of PBDE-treated products. In 

addition, releases of PBDE from sewage treatment plants (STPs) in terms of wastewater 

discharges to surface waters and the land application of sewage sludge are addressed in the life 

cycle analysis. 

2.4.1. Production Releases

The EPA’s Toxic Release Inventory (TRI) provides data on environmental releases from 

the production of decaBDE and its use as a flame retardant in the production of other materials 

for 2007 (U.S. EPA, 2009a; the latest reporting year). A total of 21.76 MT of decaBDE were 

released to the air, land, and water in 2007. Great Lakes Chemical (El Dorado, Arkansas), the 
2-12

Figure 2-2. Idealized life cycle of polybrominated diphenyl ethers if all 

information is available to complete the analysis.

major U.S. producer of decaBDE, reported a total of 16.90 MT released to the environment. 

Figure 2-3 displays a summary of fugitive, stack, surface water, and land application and 

treatment releases for 2007 (top chart on figure). DecaBDE production and manufacturing waste 

was also transferred to chemical waste landfills, bringing the total TRI reported data to 32.2 MT 

in 2007. However, the disposal of the decaBDE waste in permitted hazardous waste landfills is 

not considered as an environmental release, and, therefore, this production waste was not added

to Figure 2-3. Disposal in a permitted hazardous waste landfill is not considered to be an 

“environmental release,” because of the controls in place in the Resource Conservation and 

Recovery Act (RCRA)-permitted landfills are designed to prevent off-site migration and 

contamination of groundwater sources. Releases of decaBDE to the air from fugitive and stack 

emissions accounted for the preponderance of environmental release, ~91% of total releases, 

from the chemical manufacture of decaBDE and its use in the production of other materials. The 
2-13

Figure 2-3. TRI data showing environmental releases (kg) of decaBDE from 

production and manufacturing facilities in 2007, and total environmental 

releasesa from 1998 to 2007 (U.S. EPA, 2009a). 

Source: U.S. EPA (2009a)

a

Total environmental releases (kg/year) = fugitive air releases + stack releases + surface water releases 

+ land application disposal and treatment releases.

discharges to surface water and the off-site land treatment of the waste accounted for 3.5% and 

5.5%, respectively. The bottom chart in Figure 2-3 displays total environmental releases of 

decaBDE from production facilities from 1998 to 2007. From the chart, it appears that total 

environmental releases peaked in 1999 (at 53.9 MT) and remained essentially the same through 

1,275.12

18,545.10

766.76 1,173.44

0

2,000

4,000

6,000

8,000

10,000

12,000

14,000

16,000

18,000

20,000

Fugitvie Air Emissions Stack Air Emissions Surface Water Releases Land application disposal

Type of Environmental Release

kilograms decaBDE released 0.00E+00

1.00E+04

2.00E+04

3.00E+04

4.00E+04

5.00E+04

6.00E+04

1998 1999 2000 2001 2002 2003 2004 2005 2006 2007

kilograms per year

Releases from production 

and manufacturing in 2007

Total releases from production 

and manufacturing (1998−2007)
2-14

2002. There was a drop in environmental releases in 2003 (to 36.3 MT), followed by an increase 

to 44.8 MT in 2004. The decaBDE releases in 2007 have decreased to 1998 levels (22.7 MT in 

1998 and 21.7 MT in 2007). It is unknown if this signals a continued decrease in subsequent 

years of TRI reporting. The BDE congener distribution in decaBDE provides a basis for 

assuming the environmental loading of BDE 209, BDE 207, and BDE 208 that may be 

associated with U.S. production in 2007. Table 2-5 displays the assumed BDE congener 

distribution in the environmental releases associated with production of decaBDE in 2007. TRI 

data do not represent all releases of decaBDE from all sources and use activities, although 

substantial releases associated with industrial use are captured by the program. Therefore, the 

emissions of decaBDE to air, water, and land are likely underestimated. 

Table 2-5. Estimated environmental releases (kg) of BDE congeners from 

U.S. decaBDE production and manufacturing facilities in 2007a

Congener Stack emissions Fugitive 

emissions

Land 

application

Surface 

water

BDE 206 407.99 28.05 25.82 16.87

BDE 207 44.51 3.06 2.82 1.84

BDE 208 11.13 0.77 0.70 0.46

BDE 209 18,081.48 1,243.24 1,144.11 747.59

Total decaBDE 18,545.10 1,275.12 1,173.44 766.76

a

Calculations based on BDE congener distribution in decaBDE from Table 2-1. Releases of decaBDE are from 

TRI data as reported in U.S. EPA (2009a). 

 

2.4.2. Estimates of PBDE Releases to Indoor Air Based on Chamber Testing

PBDEs are basically flame retardant additives to materials and not bonded chemically to 

the matrix. This means that there exists a potential for the PBDEs to escape the matrix through 

the process of volatilization to the air. Thus, the products treated with PBDEs may off-gas 

PBDE within indoor air microenvironments while the product is still in use. Volatilization is one 

of the mechanisms of release from the product into the surrounding indoor air. 
2-15

Two studies were found that attempted to measure the possibility of release under 

laboratory conditions (Kemmlein et al., 2003). Kemmlein et al. (2003) determined PBDE 

emission rates by placing various products into an enclosed chamber, passing an air stream over 

the products, and systematically sampling the chamber air over the duration of the experiment. 

Test chambers consisted of volumes of either 0.02 m3 or 1 m3, and a temperature of 23oC was 

kept constant over the experiments. The products tested represented a cross-section of materials 

commonly used in interior spaces: insulation and assembly foams, IT devices, upholstered 

furniture, upholstery polyurethane foams, mattresses, and circuit boards. All materials had been 

treated with PBDE as additive flame retardants. The treated materials were kept in the enclosed 

chambers for a period of 100 days or longer in order to insure that steady-state conditions were 

reached prior to sampling. Polyurethane foam was used to passively collect PBDE vapors that 

had escaped the material matrix. Additionally chamber walls were rinsed with hexane to collect 

PBDEs that had migrated to the walls for a complete mass balance of PBDEs emitted from the 

test material. 

As part of this chamber experiment, two PC workstations (A and B) were tested in 1 m3

emission test chambers under operational conditions (Kemmlein et al., 2003). Workstation A 

consisted of a PC monitor, a desktop computer, a keyboard, a mouse, and a printer obtained from 

different manufacturers. Workstation B consisted of the same array of computing devices, but 

the components were obtained from a single manufacturer. Workstations A and B were tested in 

the air chamber for 93 and 150 days, respectively. In PC Workstation A, BDE 47, BDE 100, 

BDE 99, and BDE 85 were detected in the air at concentrations of less than 0.3 ng/m3. BDE 47, 

BDE 100, and BDE 99 were detected in the chamber air of PC Workstation B at concentrations 

of 150, 28, and 61 ng/m3

, respectively. Additionally, trace amounts of BDE 153 were detected 

in both experiments. Testing computer circuit boards in the test chamber yielded the following 

results (expressed as an emission rate in units of ng/unit-hour): BDE 17 = 0.6; BDE 28 = 1.9; 

BDE 47 = 14.2; BDE 66 = 0.4; BDE 100 = 1.3; BDE 99 = 2.6; BDE 85 = 0.1; BDE 154 = 0.1; 

and BDE 153 = 0.04. 

The PC equipment used in these tests was manufactured after 2000. To test the 

hypothesis that older plastic casings may have emitted significantly more PBDEs, Kemmlein et 

al. (2003) tested the back panel of a television casing manufactured before 1979. The back panel 

had been treated with octaBDE. BDE 28 (maximum concentration 0.5 ng/m3

), BDE 47 
2-16

(maximum concentration 8 ng/m3), BDE 66 (maximum concentration 0.24 ng/m3), BDE 100 

(maximum concentration 0.27 ng/m3), and BDE 99 (maximum concentration 0.84 ng/m3) were 

detected in the air to the test chamber. The chamber walls were rinsed with solvent, and the 

resulting solution was analyzed for the presence of PBDE congeners. BDE 47 and BDE 99 were 

detected in the rinse, corresponding to a surface concentration of 568 ng/m2

 and 514 ng/m2 of 

BDE 47 and BDE 99, respectively. It was concluded that a significant portion of BDE 47 and 

BDE 99 was adsorbed to the chamber wall; therefore, the reported air concentrations of these 

congeners were likely underestimated. From the analysis of the older plastic back panel, an 

observation can be made that older treated products may continue to emit BDEs into the air after 

long periods of time (i.e., 20 years or more). 

 A second laboratory study evaluated the potential for volatilization of PBDEs from TVs 

in Japan (Hirai et al., 2006). Three waste TV sets, obtained from a recycling facility, were 

opened, and the components were cleaned of dust and rinsed with hexane. In the experimental 

design, two TVs were turned on, and the third was left off. Separate experiments were 

conducted for each TV, with TV #1 and TV #2 kept on for a duration of 48 and 144 hours, 

respectively, and TV #3 kept in the off position. The TVs were placed in a 2.0 m (w) × 1.1 m (h) 

× 0.91 m (d) test chamber. A 1 m3

/hour air flow was introduced into the chamber, and the 

chamber was kept at a constant 28°C during the test period. Air samples were taken at the outlet 

of the chamber during the duration of the experiments. After the testing, samples were taken of 

the TV casing of TV #3. The air samples were Soxhlet extracted and then analyzed by high 

resolution gas chromatography coupled with high resolution mass spectrometry. Hirai et al.

(2006) detected total PBDEs were as follows: 9,000 ng (TV #1 kept on for 48 hours); 3,900 ng 

(TV #2 kept on for 144 hours), and 610 ng (TV #3―not turned on at all). Approximately 

76,000 mg/kg total PBDEs were detected in the TV casing. Emission factors were calculated for 

TVs #1 and #2 by dividing the total emission by PBDE content in TV casings and the length of 

test periods. From this study, emission factors (weight-to-weight emissions per year) of

4.8 × 10−6 and 7.1 × 10−7 were estimated for TVs #1 and #2, respectively. It should be noted that 

almost all of the PBDEs detected were nona- and decaBDE.

Hazratian and Harrad (2006) observed an apparent association between the ages of PCs 

used in an office and the indoor air concentrations of PBDEs. They measured PBDE levels in 

the indoor air of an office when an older and newer PC were used. There were three 
2-17

experimentally conditions: (1) Only a PC built in 1998 was used; (2) A PC built in 1998 and a 

PC built in 2003 were equally used, and (3) Only the PC built in 2003 was used by office 

employees. In Condition 1, the total PBDE concentration in indoor office air was approximately

431 pg/m3

. In Condition 2, the indoor total PBDE concentration decreased to 253 pg/m3. When 

only the newer PC was used, the total PBDE indoor air concentrations decreased further to 

81 pg/m3

. 

A study of office buildings confirmed a qualitative association between the presence of 

PBDE-treated products and subsequent levels of PBDEs in indoor air (Harrad et al., 2003). 

Indoor air with the highest PBDE levels was in rooms in office buildings equipped with 

numerous desktop PCs (12−16 per room), and numerous PUF-containing chairs (11−22 per 

room). By comparison, the lowest indoor air PBDE concentrations occurred in domestic 

environments that had no PUF-containing furniture (Harrad et al., 2003). 

PUF from upholstered furniture treated with decaBDE showed no emission of PBDE 

congeners after 168-day residence time in a test chamber (Kemmlein et al., 2003). The flexible 

polyurethane was manufactured for furniture upholstery stuffing. Thus, the investigators 

reported no detectable brominated organic compounds within the test chamber. Direct analysis 

of the material showed detectable congeners of deca- and nonaBDE in the product. 

2.4.3. Estimates of PBDE Releases to Air from Plastic Consumer Products

Attempts have been made to estimate the annual amount of PBDE that may be volatilized 

to the air during the general service life of a treated product. This is complicated by the fact that 

PBDEs are additive flame retardants rather than bonded with the plastic matrix. Two separate 

methods have been proposed for calculating the amounts of PBDEs volatilized from the product: 

an approach that utilizes the vapor pressures of the PBDEs, and an approach that calculates 

volatilization rates from the octanol-to-air partitioning coefficient. The vapor pressure approach 

was used by the EU to calculate the theoretical amount of deca-, octa-, and pentaBDE 

formulation that could have volatilized into the atmosphere over the entire European continent 

during the usage and product life of flame retardant plastic products (EU, 2001, 2002, 2003). 

Equation 2-1 was used by the EU and gives an approximate estimate of the percentage of PBDE 

that may volatilize over the product life:

 
2-18

( ) ( ) 6 = × ×× 1 1 10 Percentage loss due to volatilization . V N % p (2-1) 

where 

Vp = vapor pressure of the PBDE flame retardant, mmHg at 21°C 

N = service life of the flame retarded product in years (assumed to be 10 years by the 

EU).

Equation 2-1 was initially derived to approximate the volatilization rate of chemical 

plasticizers added to plastic films during their product use (EU, 2002). The EU concluded that 

the equation should be applicable to the estimation of PBDE loss by volatilization from a solid 

matrix because the equation emphasizes the vapor pressure of the compound as the controlling 

factor. Multiplication of the annual tonnage of PBDE-treated materials by the annual percent 

loss rate yields a rough but plausible estimate of PBDE releases to the air (kg/year) from the 

treated products. 

Assuming a vapor pressure of 3.47 ×10−8 mmHg (at 21°C) and a product life of 10 years, 

a volatilization loss rate of decaBDE from the product was calculated by the EU (using eq 2-1) to 

be 0.38% over 10 years (or 0.038% per year). Approximately 6,710 MT/year of decaBDE was 

used in plastics for all EU countries combined. Based on this quantity, the EU estimated the 

total losses of decaBDE to the air during the service life of the treated plastic products to be 

2.55 MT/year in Europe (6,710 × 0.00038; EU, 2002). Using a vapor pressure of 

4.9 ×10−8 mmHg for octaBDE, the EU estimated a volatilization rate of 0.54% over 10 years (or 

0.054% per year) (EU, 2002). Assuming a 1994 consumption figure of 2,550 MT/year, the EU 

estimated that approximately 1.38 MT/year of octaBDE volatilized from product usage 

throughout Europe (EU, 2002). PentaBDE was assumed to have a vapor pressure of 

3.5 ×10−7 mmHg, and this yielded an estimated annual loss rate of 3.9% over a 10-year product 

life (or 0.39% per year) (EU, 2001). The EU used a value of 1,100 MT/year pentaBDE in 

polyurethane foam products to calculate an annual release rate of 4.3 MT/year pentaBDE for 

Europe as a whole (EU, 2001). It should be noted that the product service life of 10 years was a 

subjective assumption by the EU, but, when combined with the average amount of PBDE-treated 

materials in use each year, is thought to give a reasonable, but unverifiable, result (Prevedouoro 

et al., 2004). 
2-19

Prevedouoro et al. (2004) estimated the volatilization flux of BDE 47 from products 

treated with pentaBDE and consumed in the United Kingdom. Using eq 2-1 and applying a high 

and low BDE 47 consumption rate (74 and 65 MT in the year 2000, respectively), Prevedouoro 

et al. (2004) calculated the annual mass flux of BDE 47 to air from the use of treated consumer 

products to be 450−750 kg in 2000. By assuming a consumption rate for each year, Prevedouoro 

et al. (2004) then estimated the annual mass flux from 1975−2000. Figure 2-4 shows these 

results. The peak in the air emissions of BDE 47 occurred in 1997 for both the high and low 

consumption values. BDE 47 emissions in 1997 were calculated to be 31 and 22.5 MT for the 

high and low consumption rates, respectively.

Figure 2-4. Volatilization of BDE 47 from PBDE-treated consumer products 

in the U.K.

Source: Prevedouoro et al. (2004). Assumes an average product life of 10 years. High and low estimates 

correspond to range of predicted consumption rates for the year 2000 (74 and 65 MT, 

respectively).

A second approach, termed the Koa approach, was used by Prevedouoro et al. (2004) to 

estimate the amount of PBDE that may volatilize from the plastic matrix. The justification for 

use of this approach was that the logarithm (base 10) of the octanol air partition coefficient (log 

Koa) of the PBDE may best describe the volatilization behavior (see Chapter 3, Section 3.2.5). 

Knowing the log Koa of the BDE congener, Prevedouoro et al. (2004) used an equation (see 

eq 2-2) developed by Breivik et al. (2002) to estimate air release emission factors for PBDE 
2-20

releases from treated plastics; Breivik et al. (2002) derived eq 2 from regressing results from 

chamber experiments involving the volatilization of polychlorinated biphenyls (PCBs) from 

commercial sealants containing PCBs. 

log EF ×−= CK +° 83.4)20(log839.0 OA (2-2) 

where 

log EF = emission factor, mass PBDE emitted/mass PBDE used per year (unitless)

Assuming a log Koa of 10.82 (at 20°C) for BDE 47, Prevedouoro et al. (2004) calculated 

an emission factor of 5.64 ×10−5. Multiplying this emission factor by the consumption rate of 

BDE 47 in 2000 (65 MT) gave an estimate of 3.5 ×10−3 MT (or 3.5 kg) BDE 47 emitted to the 

air from plastic-treated materials. The log Koa approach produced an emission estimate that was 

145 times lower than the vapor pressure approach (see eq 2-2) (Prevedouoro et al., 2004). 

Similar to eq 2-1, this Koa approach must be considered uncertain and preliminary. PBDEs are 

added to the plastic (additive flame retardants) and not covalently bonded with the plastic, 

whereas the tested sealant material used in the development of eq 2-2 likely contained PCBs that 

were bonded to the sealant material. 

These methods will be used to estimate the possible total volatilization of PBDE from 

treated plastic products in the United States. The following assumptions are used in the 

calculation: 

1. The market demand of the penta-, octa-, and decaBDE formulations for 2001 (the latest 

year for which statistics are available) is used as a surrogate for inferring the amount used 

as flame retardant for plastic products in the United States. 

2. The product life is 10 years. 

3. Two separate methods are used to calculate possible amounts of decaBDE volatilized 

from plastic: eq 2-1; the vapor pressure (Vp) method advocated by the EU, and eq 2-2 

based on the octanol air partition coefficient (log Koa). 

 

Estimates of releases to air using the EU method. In 2001, the market demand for 

decaBDE was approximately 24,500 MT in the Americas (Hites, 2004). The Americas includes 

all countries in North, Central, and South America. Roughly 80% of decaBDE is used for flame 

retardant hard plastics (Pure Strategies, Inc., 2005), with the remaining 20% used in textiles. 
2-21

Therefore, with a total demand of 24,500 MT, it is estimated that the amount of decaBDE

incorporated into new plastic products in the Americas, in 2001, is 19,600 MT. If it is assumed 

that the United States accounts for 80% of the total use of decaBDE produced in 2001, then 

15,680 MT may have been used in the United States. As discussed above, eq 2-1 produces an 

estimated volatilization loss rate of 0.038%/year―or 0.38% over 10 years. Applying these loss 

rates would result in approximately 5.99 MT decaBDE/year volatilized from these new products 

made from the flame retardant plastics. Assuming no additional products were put into the 

market during that period, this means that approximately 59.9 MT of decaBDE would have 

volatilized from plastic products over a 10-year time frame in the United States. 

The market demand for octaBDE was 1,500 MT in the Americas (Hites, 2004), or 

1,200 MT in the United States (the United States represents 80% of the total demand in North, 

South, and Central America). Using a vapor pressure of 4.9 ×10−8 mmHg for octaBDE (EU, 

2002) results in an annual estimated volatilization rate of octaBDE of 0.54% over 10 years (or 

0.054% per year). Approximately 0.648 MT/year octaBDE may volatize from plastic products 

(0.054% × 1,200 MT). 

Total market demand for pentaBDE in the Americas was estimated at 7,100 MT (Hites, 

2004), of which 95% (6,745 MT) was used as an additive flame retardant in FPUF materials 

(EU, 2001). Assuming that market demand in the United States accounted for 80% of total 

market demand in the Americas gives an estimate of 5,396 MT pentaBDE used in FPUF in the 

United States. Using eq 2-1 and a vapor pressure of 3.5 ×10−7 mmHg (EU, 2001) gives an 

estimated volatilization rate of 3.9% over a 10-year product life (or 0.39% per year). 

Multiplying the annual loss rate (0.39%/year) by the annual pentaBDE used in the United States 

(5,396 MT) produces an estimated annual volatilization of pentaBDE from FPUF of 21 MT/year. 

Table 2-6 summarizes these calculations using the vapor pressure approach (see eq 2-1). 

Estimates of volatilization flux of the individual BDE congeners contained within the 

commercial formulations were made based on their respective vapor pressures. This analysis 

represents a picture of possible annual emissions from the volatilization of PBDE commercial 

formulations and BDE congeners from treated products based on a 10-year product life. The 

analysis does not take into account the introduction of new products treated with PBDEs during 

this timeframe, but it is consistent with the approach taken by the EU in evaluating potential 

human health risks associated with exposures to commercial PBDE formulations (EU, 2003). In 
2-22

Table 2-6. Calculated PBDE emission factors and estimated PBDE emissions from plastic products in the 

United States for 2001, derived using two approaches: vapor pressure approach; log Koa approach 

Chemical

2001a,b

market 

demand 

(kg)

BDEc

content

(kg)

Vapor Pressure (Vp) Approach (see 

eq 2-1)

Log Koa Approach

(see eq 2-2)

Total BDE Congener Emissions

Summed Across Formulationsj

Vp

d 

at 20°C

(mmHg)

Vp-based 

emission 

factor

(%/yr)e

Vp-based

emissions 

(kg/yr)f

Logg

Koa

at 

20°C

Koa-based 

Emission 

factor 

(kg/kg)h

Log 

Koa-based

emissions 

(kg/yr)i

Ratio 

(Vp/ 

Koa)

BDE 

congener

Total Vp

method 

(kg)

Total Koa

method 

(kg)

DecaBDE 1.57 × 107 3.47 × 10−8 0.038 5.99 × 103 13.21 5.58 × 10−7 8.75 × 100 684 BDE 28 8.08 × 102 7.81 × 100

BDE 206 2.20% 3.45 × 105 NA NA NA NA NA NA NA BDE 47 6.14 × 103 1.33 × 102

BDE 207 0.24% 3.76 × 104 NA NA NA NA NA NA NA BDE 66 NA 3.05 × 100

BDE 208 0.06% 9.41 × 103 NA NA NA NA NA NA NA BDE 85 9.53 × 100 1.20 × 100

BDE 209 97.50% 1.53 × 107 3.47 × 10−8 0.038 5.84 × 103 13.21 5.58 × 10−7 8.53 × 100 684 BDE 99 6.00 × 102 2.41 × 101

OctaBDE 1.20 × 106 4.90 × 10−8 0.054 6.47 × 102 12.78 1.28 × 10−6 1.54 × 100 421 BDE 100 1.47 × 102 7.10 × 100

BDE 153 5% 6.00 × 104 7.46 × 10−9 0.008 4.92 × 100 11.82 8.19 × 10−6 4.91 × 10−1 10 BDE 138 NA 8.37 × 10−1

BDE 154 1% 1.20 × 104 1.98 × 10−9 0.002 2.61 × 10−1 11.92 6.75 × 10−6 8.10 × 10−2 3 BDE 153 8.24 × 101 2.86 × 100

BDE 183 40% 4.80 × 105 NA NA NA 11.96 6.25 × 10−6 3.00 × 100 NA BDE 154 2.61 × 100 8.09 × 10−8

BDE 190 1% 1.20 × 104 4.32 × 10−11 0.00005 5.70 × 10−3 NA NA NA NA BDE 183 NA 3.00 × 100

BDE 196 8% 9.60 × 104 NA NA NA NA NA NA NA BDE 190 5.70 × 10−3 NA

BDE 197 21% 2.52 × 105 NA NA NA NA NA NA NA BDE 196 NA NA

BDE 203 5% 6.00 × 104 NA NA NA NA NA NA NA BDE 197 NA NA

BDE 207 7% 8.40 × 104 NA NA NA NA NA NA NA BDE 203 NA NA

BDE 208 10% 1.20 × 105 NA NA NA NA NA NA NA BDE 207 NA NA
2-23

Table 2-6. Calculated PBDE emission factors and estimated PBDE emissions from plastic products in the 

United States for 2001, derived using two approaches: vapor pressure approach; log Koa approach (continued)

Chemical

2001a,b

market 

demand 

(kg)

BDEc

content

(kg)

Vapor Pressure (Vp) Approach (see 

eq 2-1)

Log Koa Approach

(see eq 2-2)

Total BDE Congener Emissions

Summed Across Formulationsj

Vp

d 

at 20°C

(mmHg)

Vp-based 

emission 

factor

(%/yr)e

Vp-based

emissions 

(kg/yr)f

Logg

Koa

at 

20°C

Koa-based 

emission 

factor 

(kg/kg)h

Log 

Koa-based

emissions 

(kg/yr)i

Ratio 

(Vp/ 

Koa)

BDE 

congener

Total Vp

method 

(kg)

Total Koa

method 

(kg)

PentaBDE 5.40 × 106 3.50 × 10−7 0.385 2.08 × 104 11.31 2.19 × 10−5 1.18 × 102 176 BDE 208 NA NA

BDE 28 0.2% 1.08 × 104 6.80 × 10−6 7.483 8.08 × 102 9.50 7.24 × 10−4 7.81 × 100 103 BDE 209 5.84 × 103 8.53 × 100

BDE 47 25% 1.35 × 106 4.14 × 10−7 0.455 6.14 × 103 10.53 9.89 × 10−5 1.33 × 102 46

BDE 66 1% 5.40 × 104 NA NA NA 10.82 5.65 × 10−5 3.05 × 100 NA

BDE 85 2% 1.08 × 105 8.03 × 10−9 0.009 9.53 × 100 11.66 1.11 × 10−5 1.20 × 100 8

BDE 99 35% 1.89 × 106 2.89 × 10−8 0.032 6.00 × 102 11.59 1.28 × 10−5 2.41 × 101 25

BDE 100 6% 3.24 × 105 4.13 × 10−8 0.045 1.47 × 102 11.31 2.19 × 10−5 7.10 × 100 21

BDE 138 0.5% 2.70 × 104 NA NA NA 11.13 3.10 × 10−5 8.37 × 10−1 NA

BDE 153 3% 1.62 × 105 4.35 × 10−8 0.048 7.75 × 101 11.82 8.19 × 10−6 1.32 × 100 58

BDE 154 2% 1.08 × 105 1.98 × 10−9 0.002 2.35 × 100 11.92 6.75 × 10−6 7.28 × 10−1 3

a

Market demand (kg) for the United States assuming 80% of total market demand for the Americas.

b

Percent (%) is the percent distribution of individual BDE congeners in the PBDE commercial formulation and is from Table 2-1. c

Calculated kg of BDE congener in the U.S. market demand for each PBDE commercial formulation.

d

Taken from Table 3-4 in Chapter 3: Environmental Fate of Polybrominated Diphenyl Ethers. Vp in Pascals (Pa) converted to mmHg using

1 Pa = 0.0075 mm Hg. e

Emission factor is percent released to air per year from total kg used in treatment of plastic products and is calculated using eq 2-1.

f

Derived by multiplying the Vp-based emission factor by the market demand (kg) for the PBDE commercial formulation (2nd column) or by individual BDE 

congeners contained in commercial formulation (3rd column) (kg). g

Taken from Table 3-5 in Chapter 3: Environmental Fate of Polybrominated Diphenyl Ethers.

h

Emission factor is unitless based on weight/weight (kg released to air from total kg used in treatment of plastic products) and is calculated using eq 2-2.
2-24

Table 2-6. Calculated PBDE emission factors and estimated PBDE emissions from plastic products in the 

United States for 2001, derived using two approaches: vapor pressure approach; log Koa approach (continued)

i

Derived by multiplying the Koa-based emission factor by the market demand (kg) for the PBDE commercial formulation (2nd column) or by individual BDE 

congeners contained in commercial formulation (3rd column) (kg).

j

Totals (kg) are based on summing the amounts calculated for each congener contained in each PBDE commercial formulation, i.e., penta-, octa-, and decaBDE.

NA = Data not available.
2-25

this context, the EU has suggested that the estimate of amount volatilized from products is likely 

an underestimate because the estimates are based on a single year’s product use information. 

Therefore, the EU has indicated that the actual amounts released to air could be one order of 

magnitude higher than what is predicted by eq 2-1 (EU, 2001, 2002, 2003). In order to 

compensate for this potential underestimation due to considering only 1 year of use, the EU 

increased their calculation of air releases of PBDEs from products in European countries by a 

factor of 10. A similar factor of 10 could be applied to estimates made here for the Americas, 

but given the uncertainties in the procedure (the validity of the empirical equation, the 

availability of use information before 1999, etc.) as well as the fact that penta- and octaBDE 

were taken off the market in 2004, while other new products could come into the market in 

future years, no adjustments are made here. In other words, an annual estimated loss, which is 

really 1/10 the loss over 10 full years of loss from a single year of use (2001 to be precise), will 

be assumed to generally represent any year―no additional losses are assumed based on 

introduction of new product each year. 

Estimates of releases to air using the log Koa method. From Table 2-1, it is evident that 

decaBDE is 97.5% (by weight) BDE 209. Therefore, approximately 1.53 × 107

 kg of BDE 209 

was used to treat products in the United States in 2001. The log Koa for BDE 209 has been 

calculated by Webster et al. (2006) as 13.21 (at 20°C). Then using eq 2-2, 

log EF +×−= 83.421.13839.0

log EF −= 25319.6

EF E −= 0758.5 

Multiplication of this emission factor (5.58 × 10−7) by 1.53 × 107 kg BDE 209 contained in 

decaBDE yields at estimated volatilization of 8.53 kg per year from plastic-treated products in 

the United States. Equation 2-1 predicted a volatilization rate of 5.99 MT/year (5.99 × 103

 kg). 

This same procedure was repeated for a number of BDE congeners present in the PBDE 

formulations, and the results are shown in Table 2-6. 

Table 2-6 compares the BDE emissions derived from both the EU and the log Koa

approaches (shown as the ratio of Vp/Koa). In every case, the EU approach gave a result that was 

considerably greater than that produced by the Koa approach. In general, BDE emissions 
2-26

calculated with the Vp approach are greater than the ones derived using the Koa approach, by 

upwards of 4 orders of magnitude. For example, the EU approach (see eq 2-1) produced a 

volatilized amount of BDE 209 that is 684 times greater than with the Koa approach of eq 2-2. 

Prevedouoro et al. (2004) thought that the Koa approach gave a result that was more consistent 

with atmospheric measurements of PBDEs. Harrad et al. (2003) maintains that a ratio of 

BDE 47 to BDE 99 in air of about 3−3.5 is indicative of volatilization losses from products. The 

Koa approach produced a ratio of 5.5, whereas the EU (Vp) approach resulted in a 

BDE 47:BDE 99 ratio of about 10. Hirai et al. (2006) derived decaBDE emission factors from 

measured emissions of brominated flame retardants from television sets. The televisions were 

placed in a chamber, and the air sampled for 48 or 144 hours or a wipe sample was collected. 

Hirai et al. (2006) calculated emission factors for BDE 209 of 4.8 × 10−6/tv/year (based on 

48-hour operation) or 7.1 × 10−7

/tv/year (144 hour operation). The calculated BDE 209 emission 

factor from plastics derived using the Koa approach in Table 2-6 is 5.58 × 10−7, and this falls 

within the range of emission factors developed by Hirai et al. (2006) in a chamber study. In 

consideration of these comparisons, the estimates derived from the Koa approach (see eq 2-2) are 

relied on in the life cycle analysis in this report to give plausible estimates of the amounts of 

BDE congeners volatilized from plastic products. Table 2-6 shows that the volatilization rates 

from plastic products (with the Koa approach) are 8.75 kg/year, 1.54 kg/year, and 118.30 kg/year

for deca-, octa-, and pentaBDE, respectively. From the perspective of volatilization from 

products, pentaBDE accounts for 92% of emissions of total PBDE commercial formulations, 

followed by deca- (7%) and octaBDE (1%).

 It should be noted that Hale et al. (2002) proposed the theory that the bulk degradation 

and crumbling of pentaBDE-treated polyurethane foam due to the aging of materials may give 

rise to indoor air concentrations of BDE 47, BDE 100, BDE 99, BDE 153, and BDE 154. While 

this is a possible mechanism for transferring BDE congener concentrations to indoor dust, the 

authors did not present a quantitative method for estimating the amounts of BDE congeners that 

may result in indoor air from the bulk degradation of PUF. An algorithm for estimating the 

release of PBDEs from degradation and crumbling of foam could not otherwise be found in the 

literature, and for that reason, estimates are not provided in this report for this important release 

mechanism. 
2-27

2.4.4. Estimates of the Mass Flow of PBDEs Contained in Electronic and Electrical 

Equipment (EEE) Waste in the United States

Once PBDE-treated electronic and electrical equipment (EEE) materials have reached 

their functional life, they are discarded in landfills, in incinerators, or they are recycled. This 

section describes the mass-flow analysis of environmental releases of PBDEs associated with the 

disposal of PBDE-treated EEE waste composed of products that have reached their end-life. The 

final results of the analysis described here are shown in Table 2-7. 

Table 2-7. Estimated BDE congener emissions (kg/year) to air in the United 

States from incineration of EEE waste and estimated BDE congeners in EEE 

waste that is landfilled and recycled

Incineration air 

emissions

(kg/yr)

Landfilled 

(kg/yr)

Recycled 

(kg/yr)

DecaBDE Distribution 46 857,000 153,000

BDE 206 0.022 1.012 18,854 3,366.0

BDE 207 0.0024 0.1104 2,057 367.2

BDE 208 0.0006 0.0276 514 91.8

BDE 209 0.975 44.85 835,575 149,175

OctaBDE Distribution 48 890,000 159,000

BDE 153 0.05 2.4 44,500 7,950

BDE 154 0.01 0.48 8,900 1,590

BDE 183 0.4 19.2 356,000 63,600

BDE 190 0.01 0.48 8,900 1,590

BDE 196 0.08 3.84 71,200 12,720

BDE 197 0.21 10.08 186,900 33,390

BDE 203 0.05 2.4 44,500 7,950

BDE 207 0.07 3.36 62,300 11,130

BDE 208 0.1 4.8 89,000 15,900
2-28

There is a paucity of information on the amount of PBDE-treated products that may be 

discarded in any given year. There exists some information on the amount of electronic and 

electrical equipment waste contained in municipal solid waste (MSW) in the United States (U.S. 

EPA, 2006) In the United States, an estimated 2.4 million metric tons (MMT) of EEE waste 

were generated in 2005 (U.S. EPA, 2006) and incorporated into municipal solid waste. Of this 

amount, approximately 300,000 MT (or 12.5%) of selected consumer electronics were recovered 

for recycling. Selected consumer electronics subject to recycling in the United States include 

products such as televisions, video cassette recorder (VCRs), digital video disc (DVD) players, 

video cameras, stereo systems, telephones, cell phones, hand-held electronic devices, personal 

computers, laptop computers, printers, and fax and copy machines (U.S. EPA, 2006). Most, but 

not all, of these products contain PBDE as a flame retardant (IPCS, 1994). Morf et al. (2005) 

analyzed the EEE waste at an EEE waste recycling facility in Switzerland and found an average 

concentration of 510 ±35 mg/kg and 530 ±30 mg/kg of deca- and octaBDE, respectively. In 

deriving these mean concentrations in the EEE waste, Morf et al. (2005) specifically studied 

what they termed “small” electronic waste, including, “small household appliances (e.g., toasters 

and vacuum cleaners), office and communication appliances (e.g., personal computers and 

monitors, printers, phones, and fax and photocopy machines), entertainment electronics (e.g., TV 

sets, videos, camcorders, radios, Hi-Fi stereo systems, and portable compact disk (CD players), 

and small size E&E [sic] equipment (e.g., plugs and mobile phones).” They justified this 

selection of waste as the EEE that contained the bulk of brominated flame retardant use, and 

analyzed the portions of these products most likely to contain PBDEs, such as electric circuit 

boards and TV housings. The results from their tests are shown in Tables 2-3 and 2-4 and were 

discussed earlier. Based on a mass balance of the total material input, including parts measured 

for PBDEs and parts not measured for PBDEs, they calculated total product concentrations, and 

the averages of these total product concentrations are given as 510 and 530 mg/kg for deca- and 

octaBDE, respectively. For purposes of calculating deca- and octaBDE content in EEE waste in 

the United States, the total product concentrations of deca- and octaBDE found in the Morf et al. 

(2005) study are assumed to be representative of EEE waste in the United States (as defined by 

U.S. EPA, 2006). Although imperfect, the composition of EEE waste described and analyzed by 

Morf et al. (2005) is similar enough to the composition of EEE waste in the United States 

(U.S. EPA, 2006) to permit a rough estimate of the amounts of deca- and octaBDE that may in 
2-29

present in EEE waste in the United States. Using these data, it is estimated that 1,224 MT 

decaBDE and 1,272 MT octaBDE are contained in 2.4 MMT of EEE waste generated in 2005 

(510 mg/kg deca × 2.4 MMT and 530 mg/kg octa × 2.4 MMT). The recycled portion of the EEE 

waste is estimated to contain about 153 MT decaBDE and 159 MT octaBDE (3 × 105 MT 

recycled EEE waste × 510 and 530 mg/kg for deca- and octaBDE, respectively). The amount of 

BDE congeners in the EEE waste in the United States each year is estimated assuming the low 

end of the range of distributions of BDE congeners present in the PBDE commercial formulation 

(see Table 2-1). The amounts of BDE congeners that may be present in the deca- and 

octaBDE-treated EEE waste that is recycled in the United States in 2005 are shown in Table 2-7. 

Not all of the EEE waste is recycled in the United States. The remaining 2.1 MMT of 

EEE material in MSW is disposed of either in landfills or municipal waste incinerators. Of the 

amount of total EEE waste that is not recycled, roughly 20% is combusted in MSW incinerators, 

and 80% is landfilled (U.S. EPA, 2006). This would mean that 0.42 MMT of EEE waste is 

incinerated and 1.68 MMT is sent to landfills. Using the mean concentration of deca- and 

octaBDE in EEE waste (Morf et al., 2005) that may have been incinerated or landfilled in 2005 

yields the following estimates: 

• 230 MT decaBDE in EEE waste incinerated in 2005.

• 239 MT octaBDE in EEE waste incinerated in 2005. 

• 857 MT decaBDE in EEE waste landfilled in 2005. 

• 890 MT octaBDE in EEE waste landfilled in 2005. 

BDE congeners sent to landfills: The amount of BDE congeners sent to landfills in the 

United States each year is estimated assuming the low end of the range of distributions of BDE 

congeners present in the PBDE commercial formulation (see Table 2-1). The amounts of BDE 

congeners may be present in the deca- and octaBDE-treated EEE waste that is landfilled in the 

United States in 2005 are shown in Table 2-7. 

PBDE in landfill leachate: While it is reasonable to assume that some leaching of BDE 

congeners from landfilling of treated products may occur, only limited information could be 

found in the literature to support any estimation of this as an environmental release. For

example, Osako et al. (2004) evaluated the untreated and treated leachate at seven landfills in 
2-30

Japan for the presence of BDE congeners. BDEs in the raw leachate were detected in the 

following ranges of concentration (pg/L) across the seven landfills: BDE 47 (not detected [ND]: 

2,200), BDE 28 (ND: 970), BDE 66 (ND: 3,200), BDE 99 (ND: 1,800), BDE 153 (ND: 27), and 

BDE 154 (ND; 1,200). In the treated landfill leachate, no PBDE congeners were detected, 

indicating the effectiveness of the leachate treatment process. Odusanya et al. (2009) reported on 

the PBDE distribution in the leachate from five landfills in South Africa. BDE 28, BDE 47, 

BDE 66, BDE 71, BDE 75, and BDE 77 were regularly detected in raw landfill leachate samples 

collected from all the landfill sites. No BDE 209 could be detected in any of the raw leachate 

samples. The range of concentrations of total PBDEs (pg/L) detected in each landfill was as 

follows: Landfill 1 (ND: 2,670), Landfill 2 (ND: 6,638), Landfill 3 (ND: 7,230), Landfill 4 (ND: 

41−4,009), and Landfill 5 (ND: 90−9,793). Individual congeners across all five landfills ranged 

in the following concentrations (pg/L): BDE 28 (ND: 100−3,333), BDE 47 (ND: 1,469−9,793), 

BDE 66 (ND: 4,020), BDE 71 (ND: 1,667−9,459), BDE 75 (ND: 743−7,426), BDE 77 (ND: 

4,257), BDE 85 (ND: 1,240), BDE 99 (ND: 5,191), BDE 100 (ND: 2,162), BDE 119 (ND: 

5,392), BDE 153 (ND: 875), BDE 154 (ND: 2,176), and BDE 183 (ND: 263). 

Kim et al. (2006) investigated the leaching potential of PBDEs from plastics under 

simulated landfill conditions. In the lab tests, four samples representing high-impact TV housing 

plastics were tested for leaching potential. The plastic samples were as follows: Sample 1 

consisted of HIPS containing no decaBDE; Sample 2 was HIPS TV housing with decaBDE; 

Sample 3, RC-TV, was a mixture of roughly-cut actual TV housings; and Sample 4, WC-TV, 

was well-crushed TV housings (containing decaBDE). The leaching potential was determined 

using distilled water, 20% methanol solution, and dissolved humic solution of 1,000 mg/L (based 

on organic carbon) as leachants. The leaching test conditions were a liquid-to-solid ratio of 

100:1, and a contact period of 5 days, with twice-daily agitation in a temperature-controlled 

room of 30◦

C without pH or ionic strength control. Concentrations in simulated leachate samples 

were obtained using HRGC coupled with HRMS, with the preparation of samples described by 

Kim et al. (2006). Results showed that PBDE concentrations were highest with the well-crushed 

HIPS TV housing, with total PBDE leachate concentrations of 220, 23,000, and 1,200 ng/L, 

using the leachants of distilled water, methanol, and dissolved humic acid, respectively. In all 

conditions, the nonaBDE homologue dominated all other homologue groups detected in the 

leachate samples. DecaBDE was detected in all leachate samples from the laboratory tests in the 
2-31

following ranges and test conditions: distilled water: 0.21−77 ng/L; 20% methanol: 

8.9−7,600 ng/L; dissolved humic solution: 10−210 ng/L. 

 Odusanya et al. (2009), Osako et al. (2004), and Kim et al. (2006) give an indication of 

the ranges in concentration of BDE congeners that may be present in untreated landfill leachate. 

However, these studies suggest that PBDE congeners are not expected to be detected in treated 

landfill leachate. In the United States, by federal regulation, MSW landfills are required to 

collect and treat landfill leachate, and to continuously monitor groundwater for an indication of 

leachate migration from the landfill (Code of Federal Regulations, 2008). BDE concentrations in 

treated landfill leachate are likely to be small. Due to the lack of data, no attempt was made here 

to estimate the amount of PBDE potentially released into the U.S. environment from the 

migration of landfill leachate. 

PBDE in incineration emissions: MSW incinerators in the United States have not been 

characterized for their potential stack emissions of PBDEs. PBDE congeners detected in the 

stack emissions to MSW incinerators are usually a consequence of not completely destroying 

PBDEs present in the waste during combustion (Sakai et al., 2006). MSW incinerators typically 

operate with a combustion efficiency of about 98% (Yang et al., 2007), which means that 98% of 

the PBDE content of the waste is expected to be destroyed during combustion. Previously, it 

was estimated that approximately 230 and 239 MT of deca- and octaBDE, respectively, may be 

contained in the EEE waste subject to incineration. These amounts of deca- and octaBDE 

present in the EEE waste should be destroyed by about 98%, leaving only 4.6 and 4.8 MT of 

deca- and octaBDE, respectively, subject to stack emissions. Through the use of strict 

regulations, the United States has imposed highly effective air pollution control devices on MSW 

incinerators (Federal Register, 1995). The application of dry scrubbers combined with fabric 

filters on large MSW incinerators has generally reduced the concentrations of semivolatile 

organics present in the combustion gases leaving the furnace by an additional 99% prior to 

emissions from the stack (Federal Register, 1995). The previously calculated uncontrolled 

emissions of deca- and octaBDE would be further reduced by 99% with these air pollution 

control devices, leaving only 0.046 and 0.048 MT of decaBDE and octaBDE, respectively, in the 

stack emissions from MSW incinerators in the United States (4.6 MT decaBDE × [1 − 0.99] and 

4.8 MT octaBDE × [1 − 0.99]). Table 2-7 displays the amounts of BDE congeners that may be 

emitted each year from incineration of EEE waste in the United States (assuming that the BDE 
2-32

congener distribution in Table 2-1 is not altered either by the heats of combustion or by the air 

pollution control device).

 Incineration of MSW contaminated with PBDE also forms polybrominated 

dibenzodioxins and dibenzofurans (PBDD and PBDF) in the combustion gases. PBDE is a 

direct precursor to PBDD/PBDF formation within thermal systems (Weber and Kuch, 2003). 

The mechanism for this is thought to consist of the intramolecular elimination of bromine (Br2) 

and/or hydrogen bromide (HBr) (Weber and Kuch, 2003). These debromination reactions are 

enhanced at temperatures above 500°C (Weber and Kuch, 2003), and the cleavage of the Br2 or 

HBr from the molecule leads to ring closure to form PBDD and PBDF (Ebert and Bahadir, 

2003). The kinetics of the conversion of PBDE to PBDD and PBDF appears to be more 

favorable for the lower-brominated diphenyl ethers (e.g., pentaBDE) and less favorable for the 

decaBDE (Weber and Kuch, 2003). The destruction efficiency of the PBDE contained in the 

MSW by controlled incineration ranges from 90 to 99.9%, leaving enough PBDE for the 

emission of PBDE from the stack and for the formation of PBDD and PBDF (Weber and Kuch, 

2003). Weber and Kuch (2003) and Ebert and Bahadir (2003) do not provide information on the 

efficiency of thermolytically converting the mass concentrations of PBDEs in the waste 

combusted to the mass concentrations of PBDD and PBDE formed in incinerator stack 

emissions. Therefore, no attempt is made here to estimate stack air releases of PBDDs and 

PBDFs during the incineration of MSW. 

2.4.5. Estimates of the Mass Loading of PBDEs Present in Sewage Treatment Plant 

Sewage Sludge and Effluent to Land and Water in the United States

There are approximately 16,519 STPs operating within the United States (U.S. EPA, 

2008). These STPs have a combined daily sewage flow rate of approximately 1.3 × 1011 L/day

(NAS, 2002). Approximately 93% of total daily sewage flow undergoes secondary and/or 

tertiary treatment before the wastewater is discharged into surface water (U.S. EPA, 2008). 

Applying the factor of 93% to the total daily sewage flow from all STPs combined gives an 

estimated flow rate of 1.2 × 1011 L/day for STPs with secondary and tertiary treatment. The 

secondary and tertiary treatment processes remove solids from the wastewater, concentrates the 

solids, and generates sewage sludge. Approximately 6.3 × 106 MT dwt (metric tons, dry weight) 

of sewage sludge are used or disposed of annually in the United States, of which 3.8 × 106 MT 

dwt (60%) is land applied or commercially distributed as fertilizer (NAS, 2002; U.S. EPA, 
2-33

1999). The remaining sludge is disposed of in the following manner: 1.1 ×106 MT dwt are 

landfilled (17%), 1.4 × 106 MT dwt are incinerated (22%), and 6.3 × 104 MT dwt (1%) have 

miscellaneous uses such as daily landfill cover (NAS, 2002; U.S. EPA, 1999). 

EPA recently conducted an updated national sewage sludge survey (NSSS) of PBDEs 

and other contaminants present in sewage sludges in the United States (U.S. EPA, 2009b). This 

information was used to estimate the amount of PBDE contamination in the sewage sludge that 

is applied to land as a soil amendment and fertilizer. The NSSS was a statistically based survey 

in which sewage sludges were randomly selected from 74 STPs in 35 states to represent the 

United States as a whole. Samples were collected between August 2006 and March 2007. The 

mean concentrations (μg/kg dwt) of PBDE congeners in sewage sludge were BDE 47 = 709.17, 

BDE 99 = 716.36, BDE 153 = 68.33, and BDE 209 = 2,181.23. On a national basis, the 

dominant congener was BDE 209. 

A study by North (2004) was chosen to represent the PBDE content of STP effluent 

discharges to surface waters in the United States. This is because it was a careful mass balance 

study of the distribution of PBDEs in sewage sludge and the effluent at a large tertiary STP in 

Palo Alto, California. North (2004) found that just five BDE congeners accounted for 

approximately 86−90% of total PBDEs detected in effluent and sewage sludge, respectively. 

These five congeners are BDE 47, BDE 99, BDE 153, BDE 154, and BDE 209. In the 

wastewater effluent discharged into surface water after secondary and tertiary treatment, the 

mean concentrations of the congeners were 10.5, 11.2, 0.98, 0.78, and 1.73 ng/L for BDE 47, 

BDE 99, BDE 153, BDE 154, and BDE 209, respectively. The sum of 28 BDE congeners 

detected in STP effluent at the Palo Alto STP was 29.02 ng/L (North, 2004). BDE 209 

dominated PBDE concentration in sewage sludge (at approximately 35% of the total 

concentration); however, BDE 47 and BDE 99 accounted for about 36% and 39% of the total 

concentration of the five main congeners present in treated wastewater effluent. BDE 209 

represented only 6% of the sum of the five congeners present in the STP effluent. These 

analyses suggest that sewage sludge is a major sink for BDE 209, since its concentration was 

high in sludge but not in effluent. 

Sewage sludge is also incinerated at STPs. No inventories of PBDE emissions in the 

stack gases of sewage sludge incinerators could be found in the literature. North (2004) did not 

stack test the incinerator for PBDEs at the multiple hearth sewage sludge incinerator at the Palo 
2-34 

Mass loading PBDE to land ( MT / yr ) mean concentration PBDE in sewage sludge =

x mass sludge applied to land yr (2-3) 

(1)

BDE loading MT yr to land

yr

sludgeMT

ug

kg

kg sludge

ug BDE BDE loading

47 69.2

108.310147 47 709.17

9 6

=

× × × = ×

− +

(2)

BDE loading MT yr to land

yr

sludgeMT

ug

kg

kg sludge

ug BDE BDE loading

99 72.2

108.310199 99 36.716

9 6

=

× × × = ×

− +

(3)

BDE loading MT yr to land

yr

sludgeMT

ug

kg

kg sludge

ug BDE BDE loading

153 26.0

108.310199 153 33.68

9 6

=

× × × = ×

− +

(4) BDE loading not tested in USEPA 154 = ( ) 2009

(5)

BDE loading MT yr to land

yr

sludgeMT

ug

kg

kg sludge

ug BDE BDE loading

209 29.8

108.3101209 209 23.181,2

9 6

=

× × × = ×

− +

(6)

BDE loading MT yr to land

BDE loading BDE BDE BDE BDE

96.13

47 99 153 209

Σ =

Σ ++= +

Alto STP under the assumption that PBDEs in the incoming sludge would be destroyed within 

the incineration system by >96%, leaving all PBDE congeners below the limit of detection. 

However, PBDDs and PBDFs were sampled in the incinerator emissions with the assumption 

that PBDEs entrained in the combustion gas would be thermolytically converted to 

PBDDs/PBDFs. Only one homologue group of PBDFs was detected in the emissions to the 

incinerator―triBDF, with a mean concentration of 82 μg/m3. North (2004) calculated the mass 

loading of total BDD/BDF on the basis of assuming half the detection limit for the nondetected 

homologues (hepta−octaBDD and BDF). This estimate was 2.8 × 10−7 kg/year based on total 

BDD/BDF in the stack emissions (assuming 1/2 detection limit for nondetects). 

The following mass loading calculations of BDE congeners from STPs to the 

U.S. environment are made based on mean concentrations of BDE congeners detected in the 

sludge and effluent at the Palo Alto STP. 

• Estimated annual loading of BDE congeners to the land from the land application of 

sewage sludge in the United States (mean concentrations from USEPA 2009b): 

General equation: 
2-35

• Estimated annual loading of BDE congeners to surface waters from STP effluent in 

the United States (Mean concentrations from North, 2004): 

General equation: 

Mass loading PBDE to water mean concentration P = BDE in STP effluent

x total STP effluent yr (2-4) 

(1)

BDE loading yrkg ( MT yr)to water

yr

d

d

totalL effluent

g

kg

ng

g

L effluent

ng BDE BDE loading

47 498 498.0

103.1 365

101

47 101 47 5.10

11

3

9

=

× × × × × × = ×

+

+

−

(2)

BDE loading yrkg ( MT yr)to water

yr

d

d

totalL effluent

g

kg

ng

g

L effluent

ng BDE BDE loading

99 531 531.0

103.1 365

101

99 101 99 2.11

11

3

9

=

× × × × × × = ×

+

+

−

(3)

BDE loading yrkg ( MT yr)to water

yr

d

d

totalL effluent

g

kg

ng

g

L effluent

ng BDE BDE loading

153 0465.05.46

103.1 365

101

153 101 153 98.0

11

3

9

=

× × × × × × = ×

+

+

−

(4)

BDE loading yrkg ( MT yr)to water

yr

d

d

totalL effluent

g

kg

ng

g

L effluent

ng BDE BDE loading

154 037.037

103.1 365

101

154 101 154 78.0

11

3

9

=

× × × × × × = ×

+

+

−

(5)

BDE loading yrkg ( MT yr)to water

yr

d

d

totalL effluent

g

kg

ng

g

L effluent

ng BDE BDE loading

209 0821.01.82

103.1 365

101

209 101 209 73.1

11

3

9

=

× × × × × × = ×

+

+

−

(6)

BDE loading yrkg ( ) MT yr to water

yr

d

d

totalL effluent

g

kg

ng

g

L effluent

ng BDE

BDE loading

n

n

380,1 38.1

103.1 365

101

101 02.29

28

11

3

9

28

=

× × × × × × = ×

∑

∑ ∑

=

+

+

−

=

2.5. SUMMARY OF THE GENERALIZED LIFE CYCLE ANALYSIS OF PBDEs IN 

THE UNITED STATES

This section presents a summary of a simplified and generalized life cycle analysis of 

PBDEs used in plastics, and EEE in the United States. This life cycle analysis is focused on the 
2-36

production, use, and disposal of plastics and EEE treated with deca- and octaBDE. Included in 

the life cycle is an analysis of the environmental loadings of PBDE congeners that may be 

occurring from sewage treatment plant operations in the United States. It is noted that all 

quantities generated in this example are highly uncertain. For this analysis, data and procedures 

described and summarized in previous sections are used. To the extent practical, this analysis is 

congener-specific. Figure 2-5 presents a summary of a generalized annual life cycle of PBDE in 

the United States to include production, use, disposal, recycling, and sewage treatment. All 

estimates of PBDE content of materials, EEE waste, and congener-specific emissions and 

releases are discussed and documented in Section 2.4 (above). 

Figure 2-5. Summary of a generalized annual life cycle of PBDE in the 

United States: production, use, disposal, recycling, and sewage treatment.

 The following observations are made (from Figure 2-5) within the context and limitations 

of this life cycle analysis:
2-37

1. BDE 209 is the dominant congener in every aspect of the life cycle analysis, with the 

exception of wastewater discharges from sewage treatment plant operations, in which 

case, BDE 99 is the most abundant congener; 

2. Major reservoirs for BDE 209 include municipal solid waste landfills (835,575 kg/year

landfilled) and recycled EEE waste (149,175 kg/year). The annual amount of BDE 209 

that may be released into the environment from these reservoirs is unknown and is a 

major area of uncertainty;

3. The production and manufacturing of decaBDE causes the highest amounts of BDE 209 

releases to air (19,325 kg/year) when compared to other sources analyzed, followed by 

the incineration of EEE waste (45 kg/year), and volatilization from plastic products 

(8.5 kg/year);

4. Approximately 9,434 kg/year of BDE 209 are applied to land through the application of 

industrial (1,144 kg/year) and sewage (8,290 kg/year) sludges; 

5. Production/manufacturing facilities and sewage treatment plants discharge approximately 

748 and 82 kg/year BDE 209, respectively, to surface water;

6. With regard to BDE 47, information was sufficient to estimate air releases from products 

(134 kg/year), wastewater discharges to surface water from sewage treatment plants 

(498 kg/year), and the land application of sewage sludge (2,690 kg/year);

7. Estimation of the BDE annual volatilization rates from plastic products gave the 

following rank order by amounts volatilized (high to low): BDE 47 > BDE 99 

> BDE 209 > BDE 28 > BDE 100 > BDE 66/183 > BDE 153 > BDE 85 > BDE 138/154; 

8. With regard to land-applied sewage sludge, the rank order of BDE congeners by kg/year

(high to low) is: BDE 209 > BDE 99 > BDE 47 > BDE 153; 

9. With regard to wastewater discharges from sewage treatment plant operations, the rank 

order of BDE congeners by kg/year (high to low) is BDE 99 > BDE 47 > BDE 209 

> BDE 153 > BDE 154. 

 Not all potential sources of PBDE releases to the environment were addressed in the 

generalized life cycle analysis. For example, Choi et al. (2008) found PBDE emissions 

emanating from a steel mill in South Korea using an array of air samplers near the complex. An 

average concentration of 25.2 pg of total PBDEs/m3 of air was detected (range: 9.0−61.6 pg/m3

) 

near the facility, and air dispersion modeling related these results to mill operations. Although 

indicative of steel mills as a possible source of PBDE air emissions, this study did not give a 

sufficient basis for estimating emissions from U.S. steel mills. Rather than directly measuring 
2-38

emission rates of PBDEs from the mill, the study relied on the indirect method of environmental 

monitoring and inferring impacts from these results with air dispersion analysis. Cetin and 

Odabasi (2008) also observed higher air concentrations of PBDE compounds in ambient air 

samples taken in urban areas of Turkey where steel mill operations were located. However, this 

study also did not directly measure PBDE emissions from a steel plant, and only provided 

indirect evidence that steel mills do emit PBDEs to the atmosphere. 

Other studies have implicated dust in the cabin spaces of automobiles as another source 

of PBDE releases to the environment. Lagalante et al. (2009) sampled dust from the interiors of 

60 automobiles that were available for resale at U.S. dealerships. Lagalante et al. (2009) 

detected BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209 in 

the dust. BDE 209 was the dominant congener detected, with a mean concentration of 48.1 μg/g. 

This study was in agreement with a previous study of dust obtained from 20 used automobiles in 

the United Kingdom (Harrad et al., 2008). However, Harrad et al. (2008) found a higher mean 

concentration of BDE 209 of 410 μg/g. Although these studies have implications for human 

exposures to PBDEs for time spent within the automobile, this information was judged to be 

insufficient in estimating mass flows of PBDEs to the environment and incorporation into the life 

cycle analysis. 

 Finally, the potential for PBDEs to enter the indoor (or car) environment through the 

flaking and crumbling of treated foam products cannot be characterized quantitatively, or even 

qualitatively. It is theorized to be a major source to the indoor environment, but procedures to 

estimate this kind of release from foams is not available.

Major uncertainties are associated with this generalized life cycle analysis. These include 

the following: 

• No accurate or current production figures are readily available on the amount of PBDE 

produced annually in the United States. Hence, this analysis presumed that the amounts 

corresponding to market demand are equal to the amounts produced. Moreover, it was 

assumed that 80% of the total market demand of North, Central, and South America 

(combined) is the market demand of the United States. Although this is a reasonable 

assumption when U.S. consumption patterns are considered, there is no published 

information to validate this assumption. 

• The environmental releases associated with production and manufacturing were derived 

from TRI data reported for the year 2007. These data are self-reported by industry and 
2-39

are often based on engineering calculations rather than on actual measurements. The 

accuracy of such numbers is largely unknown. 

• The estimates of air emissions of PBDEs due to volatilization from plastic products are

very uncertain. Rates of volatilization were estimated by using an equation (see eq 2-2) 

that relies on the use of the chemical’s octanol air partition coefficient to approximate the 

volatilization process. Although Prevedouoro et al. (2004) have determined that 

estimates from this method appear to be reasonable, additional research on volatilization 

losses is necessary to reduce uncertainty.

• Due to limitations in existing information, this analysis relied on the use of information 

collected over different years, i.e., not all information corresponded to the same year. For 

example, market demand figures were from 2001 (the latest available information); 

estimates of environmental releases from production and manufacturing facilities were 

for 2007; estimates of EEE waste generation amounts and amounts recycled, landfilled, 

and incinerated were for 2005, and releases from sewage treatment plant operations were 

from 2007. Combining these years into a single annual estimate, although uncertain, 

likely means that the life cycle is fairly representative of annual releases between 2001 

and 2007. 

• Although production of octa- and pentaBDE ceased by the end of 2004, estimates of 

PBDE volatilization from plastic materials assumed the presence of deca-, octa-, and 

pentaBDE in the products. The analysis presumed that plastic products were made flame

retardant in 2001 (corresponding to the market demand of PBDEs), and this predated the 

time when octa- and pentaBDE were no longer used. Because the analysis assumed a 

functional life of 10 years for the plastic products, it is logical to presume that the 

products would contain octa- and pentaBDE.

• The analysis assumed that once EEE materials were placed into landfills, there would be 

no further environmental releases. There is the potential that PBDEs may be released 

when incorporated into landfill leachate and the leachate leaves the landfill site to 

contaminate groundwater or surface water. However, strict Federal regulations are 

intended to restrict both the formation of landfill leachate and migration of leachate from 

the site. No information is currently available that directly demonstrates that PBDE 

contaminated leachate has migrated from landfills in the United States and subsequently 

contaminated surface waters or groundwater. No information is readily available on the 

amounts of leachate currently generated by U.S. landfills. 

 

2.6. SUMMARY

 This chapter has presented information on the production, uses, and life cycle of PBDE 

commercial formulations and their respective BDE congeners. It should be noted that the 

production figures are highly uncertain, and no actual amounts of PBDEs used in the United 

States for the treatment of plastics, textiles, and flexible polyethylene foam could be found. The 

industry usually aggregates data on a global scale, and statistics are available for the Americas, 
2-40

Asia, and Europe (to include the countries of the EU). The industry breakdown indicates that 

approximately 56,418 MT of PBDEs were consumed worldwide in 2003 (the latest reporting 

year) of which 83% was decaBDE. The industry voluntarily ceased production of penta- and 

octaBDE in December 2004. Today, only decaBDE is being produced in the United States, and 

the EU countries have banned the use of decaBDE effective 4/1/2008. It is estimated that 

15,680 MT decaBDE is currently consumed in the United States. Eighty percent of the 

decaBDE produced in the United States is used as an additive flame retardant to rigid plastics 

used in casings to TVs, PCs, liquid crystal display (LCD) screens, and other EEE. The 

remaining 20% are used in textiles such as fabrics used in car seats and other upholstery fabrics 

but not in clothing. 

 The current and past use of PBDE-treated products can result in environmental releases 

of PBDEs. The aim of this chapter has been to evaluate the amount of PBDEs that could 

potentially be released from the life cycle of PBDEs, including the release during production, the 

release from products in use, the release from disposal and recycling of products containing 

PBDEs, and the releases from the land application of sewage sludge and the effluent surface 

water discharges from sewage treatment plants that manage industrial and domestic sewage. 

While there are numerous data gaps in determining these quantities for all PBDEs, this chapter 

has reviewed available information on these pathways and highlighted where key information is 

lacking in an effort to comprehensively evaluate the life cycle of releases from production to use 

to end disposal. 

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Harrad, S; Wijesekera, R; Hunter, S; et al. (2003) Preliminary assessment of U.K. human dietary and inhalation 

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flame-retardant plastics. Chemosphere 65:506−513.

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Lagalante AF; Oswald, TD; Calvosa; FC. (2009) Polybrominated diphenyl ether (PBDE) levels in dust from 

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Morf, LS; Tremp, J; Gloor, R; et al. (2005) Brominated flame retardants in waste electrical and electronics 

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2-43

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2-44

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3-1

3. ENVIRONMENTAL FATE OF POLYBROMINATED DIPHENYL ETHERS

3.1. INTRODUCTION

This chapter reviews the physical and chemical properties important to understanding the 

environmental fate of polybrominated diphenyl ether (PBDEs). This chapter includes brief 

descriptions on how the chemicals move and partition among the environmental media once 

released into the open environment. The physical and chemical properties control tendencies for 

PBDEs to move into air, soil, water, and sediments, and to exchange among environmental 

compartments. They also indicate the physical form and phase of the chemical present in air and 

water. The physical and chemical properties influence the extent to which biotic and abiotic 

processes may transform or degrade PBDEs in the environment. Bioconcentration properties 

indicate the relative propensities of brominated diphenyl ether (BDE) congeners to 

bioaccumulate and biomagnify in ecological food chains. Overall, the properties most important 

for understanding the environmental behavior of PBDEs are water solubility (WS), octanol water 

partition coefficient (Kow), octanol air partition coefficient (Koa), the Henry's Law constant (H), 

and vapor pressure (Vp). 

3.2. PHYSICAL/CHEMICAL PROPERTIES

The following section is a general overview of the physical and chemical properties of 

the various PBDE congeners. Although there is the possibility of 209 BDE congeners, this 

section focuses on most of the congeners that comprise the deca-, octa-, and pentaBDE 

commercial formulations as flame retardants. 

3.2.1. Water Solubility (WS)

The chemical parameter of WS describes how readily a chemical compound (referred to 

as the solute) dissolves in water at a given temperature. This is one of the most important 

parameters in environmental chemistry. Highly soluble chemicals are easily and quickly 

distributed within surface water and groundwater. A typical method of estimating aqueous 

solubility is to add an excess amount of a pure chemical to water until equilibrium is achieved 

and the maximum concentration in water is reached. Water solubility is typically measured in 

units of milligram solute per liter of water (mg/L), and at a standard reference temperature 
3-2

(20−25oC). The higher the value, the more soluble the chemical is in water. Generally, 

chemicals are considered to have a low, medium, or high aqueous solubility if their WS is ≤0.1, 

0.1−10, and >10 mg/L (measured at 25oC), respectively (NAS, 2001). Table 3-1 summarizes the 

WS of commercial PBDE formulations and some specific PBDE congeners. PBDEs have low 

water solubility. Organic contaminants with low solubility do not readily dissolve in surface 

water (relatively hydrophobic) and typically have high log octanol water partition coefficients, 

which suggest a high absorption capacity for organic carbon in soils and sediments. 

3.2.2. Octanol Water Partition Coefficient (Kow) 

Another useful chemical property for indicating how a chemical moves between the 

aqueous phases and into sediments and biota is the octanol water partition coefficient, or Kow. 

The Kow value has become an important parameter in the prediction and understanding of the fate 

of organic chemicals in the aquatic and terrestrial environments. This property is derived 

through laboratory experiment and quantitative structure activity relationships to related 

chemicals, and it is defined as the ratio of the concentration of a contaminant in n-octanol 

(normal octanol) over the concentration of the same contaminant in water. The n-octanol is 

intended to generally represent all organic substances. As noted previously, it is related to water 

solubility in that the higher the chemical water solubility, the lower the propensity for 

bioaccumulation (the lower the value for Kow). It is expressed as a unitless value. In the 

experimental derivation of the Kow, octanol is an organic solvent used as a surrogate for organic 

matter. Although dimensionless, the Kow coefficient is usually expressed as the logarithm, 

base 10, of the ratio value (i.e., log Kow). In general, organic chemicals with an experimental log 

Kow coefficient equal to or greater than 5.0 have the property of being very hydrophobic, being 

tightly absorbed to organic matter, and possessing a tendency to bioaccumulate. Table 3-2 

summarizes log Kow values for brominated diphenyl ethers. All PBDEs tested have high log Kow

coefficients, which indicate that they have a tendency for bioaccumulation. 

3.2.3. Henry’s Law Constant (H)

Organic contaminants can transfer from water bodies into the air and from air back into 

water bodies. Henry’s Law constant (H) is an air-water partition coefficient, and it is a measure 

of the chemical’s equilibrium distribution between air and water at a specified temperature. In
3-3

Table 3-1. Estimated water solubility values for PBDEs

Brominated 

diphenyl ether

Aqueous solubility

(mg/L @ 25o C) Reference Solubility scorea

DecaBDE 

(Commercial)

<0.001 EU (2002) Low

OctaBDE 

(Commercial)

0.005

0.002

EU (2003)

ATSDR (2004)

Low

PentaBDE 

(Commercial)

0.013 EU (2001) Low

BDE 11 0.088 Palma et al. (2002) Low

BDE 15 0.13 ATSDR (2004) Low/Medium

Borderline

BDE 17 0.026 Palma et al. (2002) Low

BDE 18 0.026 Palma et al. (2002) Low

BDE 28 0.07 ATSDR (2004) Low

BDE 32 0.026 Palma et al. (2002) Low

BDE 35 0.005 Palma et al. (2002) Low

BDE 37 0.005 Palma et al. (2002) Low

BDE 39 0.005 Palma et al. (2002) Low

BDE 47 0.001−0.002 Palma et al. (2002)

ATSDR (2004)

Low

BDE 66 0.018 ATSDR (2004) Low

BDE 77 0.006 ATSDR (2004) Low

BDE 85 9 × 10−7−8 × 10−5

0.006

Palma et al. (2002)

ATSDR (2004)

Low

BDE 99 9 × 10−7−2.4 × 10−3

0.009

Palma et al. (2002)

ATSDR (2004)

Low

BDE 100 0.04 ATSDR (2004) Low

BDE 116 9 × 10−7−8 × 10−5 Palma et al. (2002) Low

BDE 119 9 × 10−7−8 × 10−5 Palma et al. (2002) Low
3-4

 

Table 3-1. Estimated water solubility values for PBDEs (continued)

Brominated 

diphenyl ether

Aqueous solubility

(mg/L @ 25o C) Reference Solubility scorea

BDE 128 4.15 × 10−6 Palma et al. (2002) Low

BDE 138 0.001 ATSDR (2004) Low

BDE 153 0.001

0.0009

ATSDR (2004)

Tittlemier et al. (2002)

Low

BDE 154 0.001

0.0009

ATSDR (2004)

Tittlemier et al. (2002)

Low

BDE 172 2.16 × 10−7 Palma et al. (2002) Low

BDE 176 2.16 × 10−7 Palma et al. (2002) Low

BDE 181 2.16 × 10−7 Palma et al. (2002) Low

BDE 183 0.002

0.0015

ATSDR (2004)

Tittlemier et al. (2002)

Low

BDE 185 2.16 × 10−7 Palma et al. (2002) Low

BDE 190 2.16 × 10−7 Palma et al. (2002) Low

BDE 209b <0.001 EU (2002) Low

a

Solubility score is as follows: low = relatively insoluble; medium = somewhat soluble; high = very soluble. b

No estimates or measurements of BDE 209 could be found in the literature. Because the decaBDE commercial 

formulation is 97% BDE-209, the value for decaBDE is assumed to represent BDE 209.

 
3-5

Table 3-2. Estimated octanol water partition coefficients (log Kow) values for 

PBDEs

Polybrominated diphenyl 

ether Log Kow coefficient Reference

DecaBDE (Commercial) 6.27

6.27

ATSDR (2004)

EU (2003)

OctaBDE (Commercial) 6.29

6.29 

ATSDR (2004)

EU (2003) 

PentaBDE (Commercial) 6.64−6.97

6.57 

ATSDR, (2004)

EU (2001) 

BDE 11 5.83 Palma et al. (2002)

BDE 15 5.86

5.74 

5.55

Kuramochi et al. (2004)

ATSDR (2004) 

Tittlemier et al. (2002)

BDE 17 5.52−5.88 Palma et al. (2002)

BDE 18 5.52−5.88

5.74

Palma et al. (2002)

ATSDR (2004)

BDE 28 5.94

5.98

ATSDR (2004)

Tittlemier et al. (2002)

BDE 32 5.52−5.88 Palma et al. (2002)

BDE 35 5.52−6.72 Palma et al. (2002)

BDE 37 5.52−6.72 Palma et al. (2002)

BDE 39 5.52−6.72 Palma et al. (2002)

BDE 47 6.01−6.77

6.81 

6.48 

6.55 

6.77

Palma et al. (2002)

ATSDR (2004) 

Kuramochi et al. (2004)

Tittlemier et al. (2002) 

Hardy (2004)

BDE 66 6.73 Tittlemier et al. (2002)

BDE 77 6.73 Tittlemier et al. (2002)

BDE 85 6.57−7.66

7.03 

Palma et al. (2002)

Tittlemier et al. (2002)
3-6

 

Table 3-2. Estimated octanol water partition coefficients (log Kow) 

values for PBDEs (continued)

Polybrominated diphenyl 

ether Log Kow coefficient Reference

BDE 99 6.53−7.66

7.32 

7.21 

7.13 

7.66

Palma et al. (2002)

ATSDR (2004) 

Kuramochi et al. (2004) 

Tittlemier et al. (2002) 

Hardy (2004)

BDE 100 7.24

6.86 

ATSDR (2004)

Tittlemier et al. (2002) 

BDE 116 6.71−7.66 Palma et al. (2002)

BDE 119 6.71−7.66 Palma et al. (2002)

BDE 128 7.39−8.55 Palma et al. (2002)

BDE 138 7.91 Tittlemier et al. (2002)

BDE 153 7.9

7.83 

7.62

ATSDR, 2004

Kuramochi et al. (2004)

Tittlemier et al. (2002)

BDE 154 7.82

7.39

ATSDR (2004)

Tittlemier et al. (2002)

BDE 172 9.44 Palma et al. (2002)

BDE 172 9.44 Palma et al. (2002)

BDE 176 9.44 Palma et al. (2002)

BDE 181 9.44 Palma et al. (2002)

BDE 183 8.27 ATSDR (2004)

BDE 185 9.44 Palma et al. (2002)

BDE 190 9.44

8.36

Palma et al. (2002)

Tittlemier et al. (2002)

BDE 209a 6.27

6.27

ATSDR (2004)

EU (2003)

a

No estimates or measurements of BDE 209 could be found in the literature. Because the decaBDE commercial 

formulation is 97% BDE-209, the value for decaBDE is assumed to represent BDE 209.
3-7

general, H is derived from the ratio of vapor pressure to the chemical’s aqueous solubility. 

Knowledge of H is essential to understanding the direction and mass flux of contaminants 

transferring from water to air. The rate of volatilization from water to air and the scavenging of 

the vapor phase of the contaminant in air by precipitation (i.e., wet deposition) are governed by 

H. Usually, H is expressed as atm-m3

/mol. Volatilization becomes an important transfer 

mechanism when the computed H is between 10−5

 and 10−3 atm-m3/mol (Ritter et al., 1995). 

Chemicals with H values greater than 10−3 atm-m3/mol rapidly volatilize into air. PBDEs are

low-volatile organic chemicals with H at 10−4 atm-m3/mol or less, with lower H at higher degrees 

of bromination. Table 3-3 summarizes estimated H for PBDEs.

3.2.4. Vapor Pressure (Vp) 

Volatilization of a chemical and its presence in air is driven by the Vp of the chemical. 

Vp also controls the phase distribution of a chemical in air (e.g., the proportion that exists in the 

vapor and particle phases). The Vp is a measure of the force per unit area (i.e., pressure) exerted 

by a chemical in vapor phase while in equilibrium with its liquid or solid phase at a specified 

temperature. Vp is usually expressed in units of Pascals (Pa) or millimeter of Hg (mmHg). In 

general, volatile organic compounds have a solid phase vapor pressure ≥10 Pa at an atmospheric 

temperature of 25oC (Olsen and Nielsen, 2001). The semivolatile organic compounds tend to 

have solid phase vapor pressures <1 Pa at 25oC. PBDEs are semivolatile organic compounds. 

Vapor pressures of PBDE decrease with increasing molecular weight and degree of bromination. 

In a theoretical context, the subcooled liquid Vps of PBDE compounds best represents their 

tendency to partition between the vapor and particle phases in air. Because of this, the scientific 

literature on chemical and physical properties has primarily reported values for PL and not for PS. 

However, PS of the BDE congeners can be calculated from the PL using eq 3-1 (Paasivirta et al., 

1999). 

log log / 1 / / 2.3026 P P S R Tm T S LF = +∆ × − ( ) ( ( ) ) (3-1) 
3-8

Table 3-3. Estimated Henry’s Law constants (H) for PBDEs

Brominated diphenyl ether (H)

(atm-m3/mol @ 25oC) Reference

DecaBDE (Commercial) 1.20 × 10−8

3.95 × 10−7

ATSDR (2004)

Cetin et al. (2005)

OctaBDE (Commercial) 7.5 × 10−8−2.6 × 10−7 ATSDR (2004)

PentaBDE (Commercial) 3.5 × 10−6−1.2 × 10−5 ATSDR (2004)

BDE 15 2.07 × 10−4

2.10 × 10−4

ATSDR (2004)

Tittlemier et al. (2002)

BDE 28 5.03 × 10−5

7.80 × 10−5

5.10 × 10−5

ATSDR (2004)

Cetin et al. (2005) 

Tittlemier et al. (2002)

BDE 47 1.48 × 10−5

8.39 × 10−6

ATSDR (2004)

Cetin et al. (2005)

BDE 66 4.93 × 10−6 ATSDR (2004)

BDE 77 1.18 × 10−5 ATSDR (2004)

BDE 85 1.09 × 10−6 ATSDR (2004)

BDE 99 2.27 × 10−6

5.92 × 10−6

ATSDR (2004)

Cetin et al. (2005)

BDE 100 6.81 × 10−7

2.37 × 10−6

ATSDR (2004)

Cetin et al. (2005)

BDE 153 6.61 × 10−7

4.34 × 10−6

ATSDR (2004)

Cetin et al. (2005)

BDE 154 2.37 × 10−6

7.90 × 10−7

ATSDR (2004)

Cetin et al. (2005)

BDE 183 7.3 × 10−8 ATSDR (2004)

BDE 209 3.95 × 10−7 Cetin et al. (2005)
3-9

where

PS = solid phase vapor pressure, Pa

PL = subcooled liquid vapor pressure, Pa

∆SF = entropy of fusion, ≈56 J k−1 mol−1

 (Wania and Dugani, 2003) 

R = ideal gas constant, 8.3143 J K−1 mol−1

Tm = melting point temperature, K 

T = reference temperature, K

Table 3-4 summarizes the calculated solid phase (PS) and subcooled liquid vapor pressures (PL) 

of some BDE congeners. 

3.2.5. Octanol Air Partition Coefficient (Koa) 

 The octanol air partition coefficient, log Koa, is the ratio of the concentration of the 

chemical in air versus the concentration of the chemical in octanol when the octanol air system is 

at equilibrium (Harner and Shoeib, 2002). The value of Koa is dimensionless. As with Vp, log 

Koa is temperature dependent. For every 10oC decrease in atmospheric temperature, there is a 

corresponding two- to threefold increase in partitioning of semivolatile organics to the organic 

phase (Harner and Shoeib, 2002). In broad terms, the log Koa is suggestive of the environmental 

cycling of semivolatile compounds between the air and organic phases such as soil particles, air 

particles, and vegetation. The underlying assumption of the log Koa is that n-octanol is a good 

surrogate for absorption of all organic carbon. The greater the Log Koa (e.g., ≈10) is, the stronger 

the propensity of the atmospheric BDE congener to absorb to the organic content of soils and 

vegetation (Wania et al., 2002). It is, therefore, a relative indicator of chemical mobility, and the 

tendency to exchange from the atmosphere to the surface. Log Koa values from 6−11 indicate 

that atmospheric PBDEs strongly absorb into forest canopies and other vegetative biomass 

(Su et al., 2007). Table 3-5 summarizes the octanol air partition coefficients (log Koa) of PBDE 

compounds. 
3-10

Table 3-4. Estimated solid phase vapor pressures (PS) and subcooled liquid vapor pressures (PL) of some PBDEs 

(Pascals at 25°C)

Brominated diphenyl ether

T

m 

(K)

T

m 

Reference PS

a PL

PL

Reference

DecaBDE (Commercial) 573.15 EU (2002) 9.28 × 10−9 4.63 × 10−6 EU (2002)

OctaBDE (Commercial) 473.15 EU (2003) 1.26 × 10−7 6.59 × 10−6 EU (2003)

PentaBDE (Commercial) 475.15 EU (2002) 8.60 × 10−7 4.69 × 10−5 EU (2001)

BDE 1 NA 0.163 Wong et al. (2001)

BDE 2 NA 0.128 Wong et al. (2001)

BDE 3 NA 2.59 × 10−1 Tittlemier et al. (2002)

BDE 7 NA 1.68 × 10−2 Wong et al. (2001)

BDE 8 NA 1.37 × 10−02 Wong et al. (2001)

BDE 10 NA 2.77 × 10−2 Wong et al. (2001)

BDE 12 NA 1.19 × 10−2 Wong et al. (2001)

BDE 15 329.15 Marsh et al. (1999) 8.59 × 10−3

4.88 × 10−3

1.73 × 10−2

9.84 × 10−3

2.00 × 10−2

Tittlemier et al. (2002)

Wong et al. (2001) 

Wania et al. (2002)

BDE 28 337.15 Marsh et al. (1999) 9.07 × 10−4

6.51 × 10−4

2.19 × 10−3

1.57 × 10−3

Tittlemier et al. (2002)

Wania et al. (2002)

BDE 30 358.15 Marsh et al. (1999) 1.18 × 10−3 4.56 × 10−3 Wong et al. (2001)

BDE 32 350.40 Palma et al. (2002) 6.91 × 10−4 2.25 × 10−3 Wong et al. (2001)
3-11

 

Table 3-4. Estimated solid phase vapor pressures (PS) and subcooled liquid vapor pressures (PL) of some 

PBDEs (Pascals at 25°C) (continued)

Brominated diphenyl ether

T

m 

(K)

T

m 

Reference PS

a PL

PL

Reference

BDE 33 NA 1.49 × 10−3 1.49 × 10−3 Wong et al. (2001)

BDE 35 413.01 Palma et al. (2002) 1.04 × 10−4 1.39 × 10−3 Wong et al. (2001)

BDE 37 321.65 Palma et al. (2002) 6.00 × 10−4 1.02 × 10−3 Wong et al. (2001)

BDE 47 353.65 Palma et al. (2002) 7.42 × 10−5

5.52 × 10−5

2.50 × 10−4

1.86 × 10−4

2.15 × 10−4

Wong et al. (2001)

Tittlemier et al. (2002) 

Wania et al. (2002)

BDE 66 NA 2.38 × 10−4

1.22 × 10−4

Wong et al. (2001)

Tittlemier et al. (2002) 

BDE 69 NA 4.00 × 10−4 Wong et al. (2001)

BDE 75 408.15 Marsh et al. (1999) 4.10 × 10−5 4.92 × 10−4 Wong et al. (2001)

BDE 77 368.15 Marsh et al. (1999) 3.21 × 10−5

1.40 × 10−5

1.56 × 10−4

6.79 × 10−5

Wong et al. (2001)

Tittlemier et al. (2002)

BDE 82 NA 4.80 × 10−5 6.47 × 10−5 Wong et al. (2001)

BDE 85 396.45 Palma et al. (2002) 1.07 × 10−6 9.86 × 10−6 Tittlemier et al. (2002)

BDE 99 365.45 Palma et al. (2002) 1.49 × 10−5

3.85 × 10−6

7.94 × 10−6

6.82 × 10−5

1.76 × 10−5

3.63 × 10−5

Wong et al. (2001)

Tittlemier et al. (2002) 

Wania et al. (2002)
3-12

Table 3-4. Estimated solid phase vapor pressures (PS) and subcooled liquid vapor pressures (PL) of some 

PBDEs (Pascals at 25°C) (continued)

Brominated diphenyl ether

T

m 

(K)

T

m 

Reference PS

a PL

PL

Reference

BDE 100 371.15 Marsh et al. (1999) 5.50 × 10−6

7.07 × 10−6

2.86 × 10−5

3.68 × 10−5

Tittlemier et al. (2002)

Wania et al. (2002)

BDE 115 NA 3.02 × 10−5 Wong et al. (2001)

BDE 138 NA 1.58 × 10−6 Tittlemier et al. (2002)

BDE 153 456.15 Palma et al. (2002) 1.63 × 10−7 

5.80 × 10−6

5.80 × 10−6

2.09 × 10−6

8.87 × 10−6

Wong et al. (2001)

Tittlemier et al. (2002) 

Wania et al. (2002)

BDE 154 416.15 Marsh et al. (1999) 2.64 × 10−7 3.80 × 10−6 Tittlemier et al. (2002)

BDE 183 NA 4.68 × 10−7 Tittlemier et al. (2002)

BDE 190 470.4 Palma et al. (2002) 1.85 × 10−8 

5.76 × 10−9

9.05 × 10−7

2.82 × 10−7

Wong et al. (2001)

Tittlemier et al. (2002)

BDE 209b 573.15 EU (2002) 9.28 × 10−9 4.63× 10−6 EU (2002)

a

PS was calculated from PL using eq 3-1 and assuming the melting point temperature (Tm) in the table. NA = not available in the literature.

b

No estimates or measurements of BDE 209 could be found in the literature. Because the decaBDE commercial formulation is 97% BDE 209, the value for 

decaBDE is assumed to represent BDE 209.
3-13

Table 3-5. Estimated octanol air partition coefficients (log Koa) of PBDEs

Brominated diphenyl ether Log Koa 

(@ 25oC) Reference

BDE 1 7.24 Wania et al. (2002)

BDE 2 7.36 Wania et al. (2002)

BDE 7 8.37 Wania et al. (2002)

BDE 8 8.47 Wania et al. (2002)

BDE 10 8.12 Wania et al. (2002)

BDE 12 8.55 Wania et al. (2002)

BDE 13 8.57 Wania et al. (2002)

BDE 15 8.64 Wania et al. (2002)

BDE 17 9.30 Harner and Shoeib (2002)

BDE 21 9.49 Wania et al. (2002)

BDE 28 9.50 Harner and Shoeib (2002)

BDE 30 9.02 Wania et al. (2002)

BDE 32 9.28 Wania et al. (2002)

BDE 35 9.48 Wania et al. (2002)

BDE 37 9.68 Wania et al. (2002)

BDE 47 10.53

10.34

Harner and Shoeib (2002)

Wania et al. (2002)

BDE 66 10.82

10.49

Harner and Shoeib (2002)

Wania et al. (2002)

BDE 69 10.23 Wania et al. (2002)

BDE 75 10.13 Wania et al. (2002)

BDE 77 10.87

10.70

Harner and Shoeib (2002)

Wania et al. (2002)

BDE 82 11.14 Wania et al. (2002)

BDE 84 11.52 Wania et al. (2002)

BDE 85 11.66 Harner and Shoeib (2002)

BDE 99 11.31

11.28

Harner and Shoeib (2002)

Wania et al. (2002)

BDE 100 11.13 Harner and Shoeib (2002)

BDE 126 11.97 Harner and Shoeib (2002)

BDE 153 11.82

12.15

Harner and Shoeib (2002)

Wania et al. (2002)

BDE 154 11.92 Harner and Shoeib (2002)

BDE 156 11.97 Harner and Shoeib (2002)

BDE 183 11.96 Harner and Shoeib (2002)

BDE 209 13.21 Webster et al. (2006)
3-14

3.2.6. Vapor Particle Partitioning in Air

Semivolatile organic compounds are present in the ambient air partition between the

vapor phase and the particle phase. The physics of this behavior is controlled by the Vp of the 

chemical, the temperature of the surrounding air, and the availability of airborne particulate 

matter. In terms of environmental fate, the importance of this phenomenon is that the 

aerosol-bound portion of the contaminant is subject to transport through the atmosphere over 

large geographical distances. The wet and dry deposition of the contaminated particles is the 

most significant mechanism for removal of the particle phase of PBDEs from the air. The vapor 

phase is a good predictor of the air-to-leaf transfer of the semivolatile organic compound and of 

the possibility for deposition into leafy vegetation, which is integral to the terrestrial food chain. 

Thus, the vapor phase portion of the contaminant may be more significant in terms of human 

exposures by way of the dietary pathway. 

The fraction of the semivolatile compound that is particle bound (Ф) can be calculated 

from the subcooled liquid Vp using the Junge-Pankow model as indicated in eq 3-2 (Su et al., 

2006). 

( ) ( ) /( ) L Φ= Θ + Θ c Pc (3-2) 

where 

Ф = fraction of the compound adsorbed to aerosol particles 

PL = saturation subcooled liquid phase Vp, Pa 

Θ = the particle surface area per unit volume of air, cm2 aerosol/cm3 of air 

c = a constant, 17.2 Pa-cm. 

Table 3-6 provides the particle and vapor phases of the PBDE commercial formulations and 

BDE congeners calculated from eq 3-2 and assuming the subcooled liquid Vp in Table 3-4 and an 

ambient air temperature of 25°C. The value for Θ in the equation assumes the particle surface 

area per unit volume air of 1.5 × 10−6 cm2

/cm3 typical of aerosols in background urban air 

(Whitby, 1978). The predicted particle phase of the PBDE increases with decreasing Vp and 

increasing number of bromine atoms on the molecule. 
3-15

Table 3-6. Calculated theoretical vapor particle partitioning of PBDE 

congeners in ambient air at 25°C (calculated using eq 3-2)

Brominated diphenyl 

ether congener

Brominated diphenyl

homologue Vapor phase, % Particle phase, %

BDE 1 monoBDE 100 0

BDE 2 monoBDE 100 0

BDE 3 monoBDE 100 0

BDE 7 diBDE 100 0

BDE 8 diBDE 100 0

BDE 10 diBDE 100 0

BDE 12 diBDE 100 0

BDE 15 diBDE 100 0

BDE 28 triBDE 99 1

BDE 30 triBDE 99 1

BDE 32 triBDE 99 1

BDE 33 triBDE 98 2

BDE 35 triBDE 98 2

BDE 37 triBDE 98 2

BDE 47 tetraBDE 90 10

BDE 66 tetraBDE 87 13

BDE 69 tetraBDE 94 6

BDE 75 tetraBDE 95 5

BDE 77 tetraBDE 81 19

BDE 82 pentaBDE 71 29

BDE 85 pentaBDE 28 72

BDE 99 pentaBDE 61 39

BDE 100 pentaBDE 56 44

BDE 115 pentaBDE 54 46

BDE 138 hexaBDE 6 94

BDE 153 hexaBDE 7 93

BDE 154 hexaBDE 13 87

BDE 183 heptaBDE 2 98

BDE 190 heptaBDE 1 99

BDE 209 decaBDE 1 99

 

By these calculations, and at the air temperature of 25°C, greater than 98% to 100% of 

the mono-, di-, triBDE congeners may be present in the vapor phase in ambient air. As the 

degree of bromination increases, the BDE congeners increasingly partition to the particle phase. 
3-16

For example, the various tetra- and pentaBDE congeners begin to distribute more to atmospheric 

particles than the lesser brominated compounds. The hexa- and hepta-congeners may be present 

in the particle phase from 87 to 99% at an ambient air temperature of 25°C. Approximately 99% 

of BDE 209 is expected to be associated with airborne particles. As seen in Table 3-6, 72% of 

BDE 85 is particle bound, and this appears to be an anomalous calculation due to the fact that 

Tittlemier et al. (2002) may have overestimated the subcooled liquid vapor pressure at 25°C. 

About 50−70% of pentaBDE congeners are predicted be present in the vapor phase. These 

predicted BDE phase distributions in ambient air agree with the observations of Strandberg et al. 

(2001) for distributions in air over the Great Lakes. Strandberg et al. (2001) found that at 20°C, 

about 80% of the tetrabromo homologues are in the vapor phase and about 70% of the 

hexabromo homologues are associated with the particle phase. In a study by Chen et al. (2006), 

the estimated phase distribution of PBDE congeners measured in urban air of southern China 

was as follows: BDE 28 (97% vapor phase); BDE 47 (80% vapor phase); BDE 66 (77% vapor 

phase); and BDE 99 through BDE 154 (ranged from 50−85% in the particle phase). No ambient 

air temperature was reported. 

3.3. BIOACCUMULATION, BIOCONCENTRATION, AND BIOMAGNIFICATION 

OF PBDEs IN THE AQUATIC ENVIRONMENT

 Bioaccumulation describes a process whereby an organism acquires a body burden of a 

chemical in relation to contact through all possible pathways of exposure (i.e., dietary 

absorption, transport across the respiratory surface, dermal absorption, and inhalation; Gobas and 

Morrison, 2000). Bioaccumulation for an aquatic organism occurs from contact of the organism 

with a chemical contaminant in the water column, the sediments, and through the organism’s 

food chain. The parameter that is often used to model this entire process is the bioaccumulation 

factor (log BAF). This contrasts the bioconcentration factor (BCF), which has been used to 

measure the accumulation from the water column only. In the case of lipophilic and 

hydrophobic chemicals, bioaccumulation into fish becomes important in terms of human fish 

consumption. The log BAF is derived as a logarithm of the ratio of the concentration of a 

chemical in the tissue of an aquatic organism over the concentration of the chemical in water, in 

a real aquatic setting. A laboratory-derived BCF entails an experimental setup where the 

accumulation from the water column only is determined from the concentration of the 

contaminant in the fish tissue divided by the concentration in the water. The log BAF is 
3-17

expressed in units of liters per kilogram of tissue and can be normalized to apply to wet weight 

(wwt), dry weight (dwt), or percent lipid of the organism. The biomagnification factor (BMF) is 

the ratio of the concentration of the chemical in an organism to the concentration of the chemical 

in the diet of the organism. The BMF is an indication of increases in concentration of the 

chemical as it moves up trophic levels in the aquatic ecosystem. 

Only a few studies could be located in the scientific literature that have estimated BAFs 

for PBDE congeners. Table 3-7 summarizes BAF and BMF for various aquatic species.

 The United Nations Environmental Program (UNEP) established a screening level for 

assuming high potential for bioaccumulation of the contaminants into aquatic species (WWF, 

2005). The criterion for a high potential for bioaccumulation is that the BCF or BAF in aquatic 

species for the chemical is greater than 5,000 (log BAF = 3.7). With the exception of BDE 85, it 

appears that the tri-, tetra-, and pentaBDE congeners exceed these criteria, and, therefore, have a 

potential for bioaccumulating in aquatic organisms. BDE 209 has a very low potential for 

bioaccumulating within the aquatic food web. 

3.4. BIOTIC AND ABIOTIC DEBROMINATION AND TRANSFORMATION OF 

PBDEs 

Certain biotic and abiotic processes can transform PBDEs in the environment. The 

processes most important to the breakdown of PBDEs include biodegradation, biotransformation, 

and photolysis. Biodegradation involves the breakdown of PBDEs by aerobic and anaerobic 

microorganisms into smaller compounds. The microbial organisms transform the contaminants 

through metabolic or enzymatic processes. Biotransformation is the conversion of the chemical 

structure of the PBDEs through metabolic pathways. Similar to biodegradation, the reaction is 

catalyzed by enzymes, but the process occurs in vivo in animals. Photolysis involves the 

breakdown of PBDEs by the action and the energy of sunlight. All of these environmental fate 

processes can involve the stripping of bromine atoms from the molecule, a process referred to as 

‘debromination’. With debromination, the higher-brominated BDE congeners can breakdown to 

form lower-brominated species. This section will focus on each of these degradation pathways. 

It should be stressed that there is only limited scientific information supporting the action of 

these pathways in degrading, transforming, and debrominating PBDEs. 
3-18

Table 3-7. Estimated BAF and BMF values for various aquatic species

Brominated 

diphenyl ether Species Factor Reference

Bioaccumulation factors (BAFs)

BDE 28 Lake trout 7.6 Tomy et al. (2004)

BDE 47 Lake trout

Blue mussels

7.3

6.1

Streets et al. (2006)

Gustafsson et al. (1999)

BDE 66 Lake trout 7.3 Streets et al. (2006)

BDE 85 Lake trout 2.3 Tomy et al. (2004)

BDE 99 Lake trout

Blue mussels

6.7

6.1

Streets et al. (2006)

Gustafsson et al. (1999)

BDE 10 Lake trout 7.5 Streets et al. (2006)

Biomagnification factors (BMFs)

BDE 15 Porpoise 1.6−2.4 Ramu et al. (2006)

BDE 28 Lake trout

Porpoise

7.6

1.6−2.4

Tomy et al. (2004)

Ramu et al. (2006)

BDE 47 Lake trout

Coho salmon 

Porpoise

2.1 

3.2 

1.5

Tomy et al. (2004)

Stapleton and Baker (2003) 

Ramu et al. (2006)

BDE 66 Lake trout 7.8 Tomy et al. (2004)

BDE 99 Lake trout

Porpoise

6.6 

1.8−2.4

Tomy et al. (2004)

Ramu et al. (2006)

BDE 100 Lake trout

Porpoise

6.5 

1.7−2.4

Tomy et al. (2004)

Ramu et al. (2006)

BDE 138 Lake trout 3.2−8.7 Tomy et al. (2004)

BDE 153 Lake trout

Coho salmon 

Porpoise

9.4 

4.0 

1.6−2.2

Tomy et al. (2004)

Stapleton and Baker (2003) 

Ramu et al. (2006)

BDE 154 Lake trout

Porpoise 

13.3 

1.4−2.2

Tomy et al. (2004)

Ramu et al. (2006) 

BDE 183 Lake trout

Porpoise

3.9 

0.8−2.2

Tomy et al. (2004)

Ramu et al. (2006)

BDE 190 Lake trout 1.6−5.1 Tomy et al. (2004)

BDE 209 Lake trout 0.3 Tomy et al. (2004)

 
3-19

3.4.1. Microbial Degradation of PBDEs

In theory, microbial communities can degrade and transform organic contaminants 

present in soils and sediments. This occurs because the microorganisms use the contaminants as 

a source of carbon (NAS, 1993). Carbon is a building block in the development of new cells 

during reproduction and growth. In addition, the microbes can extract electrons from the organic 

contaminant to obtain energy. The microorganism gains energy by breaking chemical bonds and 

transferring electrons away from the contaminant (NAS, 1993).

One variation of microbial degradation of organic compounds is reductive 

dehalogenation. In this process, the microbes catalyze reactions that promote the replacement of 

a halogen atom with a hydrogen atom on the organic compound. This is the primary microbial 

degradation pathway for higher molecular weight PBDEs. Although the higher brominated 

congeners are debrominated aerobically, the lower molecular weight congeners that are products 

of this process may be further debrominated by aerobic bacterial degradation via oxidative 

dehalogenation (Kim et al., 2007). 

Several recent studies have provided evidence of microbial mediated reductive 

debromination of decaBDE and octaBDE under laboratory conditions. Research is currently 

focused on anaerobic microbial strains that have demonstrated the ability to dehalogenate 

organic compounds in vitro. In particular, Dehalococcoides ethenogenes 195, Dehalococcoides 

sp. strain BAV1, and Sulfurospirillum multivorans can dechlorinate a broad range of chlorinated 

compounds including trichloroethylene (TCE), perchloroethylene, chlorobenzenes, 

polychlorinated biphenyls (PCBs), and polychlorinated dibenzo-p-dioxins (Fennell et al., 2004; 

Wu et al., 2002), and, therefore, are being investigated for their potential to debrominate PBDEs 

(He et al., 2006). 

Sulfurospirillum multivorans bacterium incubated with decaBDE has induced reductive 

debromination of decaBDE in vitro to yield octaBDE and heptaBDE after a contact time of 

2 months (He et al., 2006). The octa- and heptaBDE did not further debrominate in the presence 

of the microbe―even after 1 year. Thus, the microorganism was specific to the degradation of 

decaBDE and is incapable of debrominating lower-brominated PBDE compounds. The 

organism normally has an affinity for TCE, but when decaBDE was dissolved in TCE and 

exposed to S. multivaornas, reductive debromination did not occur. However, TCE was 

completely dechlorinated to cis-DCE. 
3-20

Dehalococcoides ethenogenes 195 bacterium was experimentally tried in an attempt to 

debrominate octaBDE (He et al., 2006). D. ethenogenes of the 195 strain contains reductive 

dehalogenase genes and is the only bacterium known to reductively dechlorinate 

tetrachloroethene and trichloroethane to ethane (Seshadri et al., 2005). D. ethenogenes

strain 195 was initially isolated from an anaerobic sewage treatment plant (STP) digester 

containing sewage sludge (Seshadri et al., 2005). When incubated for 6 months with octaBDE 

dissolved in TCE, the bacterium showed marked debromination of octaBDE to yield penta, hexa, 

and heptaBDE congeners (He et al., 2006). However, when incubated with octaBDE without the 

TCE solvent, no debromination occurred. The authors combined a number of Dehalococcoides 

species in a product they called ANAS195 and incubated it with octaBDE dissolved in TCE for 

12 months. One hundred thirty nmoles octaBDE composed of BDE 153, BDE 183, BDE 196, 

BDE 203, BDE 207, and BDE 208 added to the culture yielded 11.5 nmoles of combined 

BDE 47, BDE 49, BDE 99, and BDE 154. 

When normalized on a cell-count basis, the ANAS195 culture had a rate of 

debromination of octaBDE that was twice the rate of D. ethenogenes 195. An additional strain 

of Dehalococcoides species, Dehalococcoides sp. strain BAV1, is the only known 

microorganism that dechlorinates the lower-chlorinated organic compounds (e.g., 

dichloroethane, to ethane; Krajmalnik-Brown et al., 2004), and, therefore, He et al. (2006) 

investigated this species for the potential to further debrominate tetra- and pentaBDE congeners. 

The experiment showed that bromines in the 2 (ortho) and 4 (para) positions on the PBDE 

molecule (e.g., BDE 99 and BDE 47) were the most resistant to microbial reductive 

debromination. 

In another study, Gerecke et al. (2005) debrominated BDE 209 in vitro to yield BDE 206, 

BDE 207, and BDE 208 by contact with anaerobic mesophilic microorganisms indigenous to raw 

sewage sludge (microbial species not identified). Methane was formed as a product of microbial 

respiration, and the amount of BDE 209 decreased by 30% within 238 days. This disappearance 

rate of BDE 209 was statistically significant (p = 0.037), and it corresponded to a pseudo 

first-order degradation rate constant of 1 × 10−3

 per day, which is equivalent to a half-life of 

approximately 690 days. Debromination of BDE 209 was evident in the subsequent formation of 

nonaBDE and a number of unresolved octaBDE congeners. The removal of bromine atoms in 

the para and meta positions on the BDE 209 molecule resulted in the formation of BDE 208 and 
3-21

BDE 207, respectively. These products were formed at concentrations that were 5% of the initial 

concentration of BDE 209. Although BDE 206 was observed in the experiment, it could not be 

concluded unambiguously that it was a product of microbial degradation of BDE 209. 

Tokarz et al. (2008) observed reductive debromination of BDE 209 from laboratory

experiments simulating anaerobic conditions of sediment. Precleaned pond sediments were 

mixed with BDE 209, BDE 99, and BDE 47 to yield a final concentration of 5.0 μg/g of the BDE 

congeners in sediment. The sediment was transferred to 125-mL serum bottles, and dextrose and 

methanol were added to promote the rapid onset of anaerobic and methanogenic conditions. The 

sediments were placed into contact with anaerobic microorganisms (species not specified) and 

allowed to incubate. Extracts were prepared and analyzed with gas chromatography coupled 

with an electron capture detector and gas chromatography coupled with a mass spectrometer. In 

a second experiment, termed a ‘biomimetic’ experiment, 0.03 millimoles BDE 209, BDE 99, or 

BDE 47 were mixed with 5.0 millimoles titanium citrate and 0.2 millimoles Vitamin B12 in 

0.33-moles buffer solution. The final total volume was 31 mL. The authors noted that titanium 

citrate is an electron-donating reducing agent. Reductive debromination was enhanced in the 

biomimetic experiment; BDE 209 completely debrominated within 24 hours, forming 

lower-brominated congeners BDE 47, BDE 66, BDE 99, and BDE 119. In the anaerobic 

sediment microcosm experiments, BDE 209 more slowly debrominated through reductive 

debromination. The authors calculated that the degradation half-life of BDE 209 in the live 

microcosms ranged from 6 to 50 years, with an average of just over 14 years. 

Rayne et al. (2003a) demonstrated the in vitro anaerobic microbial reductive 

debromination of BDE 15. A nondifferentiated anaerobic bacterium from the in situ remediation 

of contaminants in a polluted river system was colonized by passing contaminated water through 

a bioreactor. Reductive debromination of BDE 15 occurred in vitro at hydraulic retention times 

of 3.4 and 6.8 hours in a fixed-film, plug-flow biological reactor. The products formed were 

BDE 3 (4-monoBDE) and nonsubstituted diphenyl ether. 

3.4.2. In Vivo Metabolic Debromination in Animals

Once absorbed into an organism, metabolic processes can break down PBDEs. There is a 

growing body of scientific evidence to suggest that certain freshwater and marine fish species 
3-22

and marine mammals are capable of metabolically debrominating PBDE congeners in vivo. This 

section is an overview of available information on debromination in wildlife. 

3.4.2.1. Evidence for Debromination in Fish

Certain fish species have the capacity to debrominate PBDE congeners in vivo. This 

involves the removal of bromine atoms in the para and meta positions of a higher molecular 

weight PBDE congener to form lower-brominated PBDE compounds. Debromination appears to 

mainly transpire in the gastrointestinal tract and liver, and by one of three possible pathways: the 

deiodinase thyroid hormone-regulating enzymes (Stapleton et al., 2006a; Tomy et al., 2004), 

endogenous microbial activity in the gut (Stapleton et al., 2004), or the enzymes of the 

microsomal mono-oxygenase system (Stapleton et al., 2004). It is not known whether all fish 

have the metabolic capacity to biotransform PBDEs, nor is it known whether the rate of 

debromination of PBDEs varies among fish species.

Laboratory studies of rainbow trout, lake trout, and carp, involving fish food spiked with 

pure BDE 209, have clearly shown accumulation of lower-brominated BDE congeners not 

initially present in the feed. This evidence is suggestive of metabolic synthesis of 

lower-brominated congeners through debromination of BDE 209 (Stapleton et al., 2004, 

2006a, b; Tomy et al., 2004; Kierkegaard et al., 1999). 

The possibility that metabolic debromination of decaBDE may be occurring in fish was 

initially observed in the late 1990s in an in vivo rainbow trout study (Kierkegaard et al., 1999). 

In this study, rainbow trout were fed fish food spiked with 7.5−10 mg of decaBDE/kg of body 

weight/day for 16, 49, or 120 days. Muscle tissues and livers from exposed fish and controls 

were analyzed separately for PBDE. Preferential accumulation of PBDE in the liver was 

observed over the 120-day study period. Due to a lack of pure analytical standards for the BDE 

congeners, it was not possible to identify specific BDE congener present in the fish; only 

analysis of homologue groups was possible. However, hexaBDE and the first eluting octaBDE 

congeners were identified as possible products of metabolic debromination of decaBDE, because 

these congeners were absent in both the spiked fish food and the control fish. Early suggestion 

of the potential for debromination of PBDEs to occur in fish prompted further in vivo studies.

Tomy et al. (2004) investigated the capacity of juvenile lake trout to debrominate a broad 

range of BDE congeners. The fish were exposed in aquaria to known amounts of 13 BDE 
3-23

congeners spiked onto fish food. The BDE congeners included BDE 28, BDE 47, BDE 66, 

BDE 77, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, BDE 190, and 

BDE 209. The duration of dietary exposure to the spiked food was for 56 days, followed by 

112 days of depuration. Five fish were sacrificed, and tissues were sampled on Days 0, 7, 14, 28, 

and 56 of the period of feeding and on Days 7, 14, 28, 56, and 112 of the depuration period. The 

strongest evidence from this study for the debromination of higher-molecular weight BDE 

congeners in lake trout comes from the comparison of patterns of BDE congeners present in 

untreated fish tissues to congeners present in the spiked fish food. The presence of one penta (an 

unidentified congener) and two hexaBDEs (BDE 140 and an unidentified hexaBDE) in the fish 

tissues served as an indicator of the occurrence of in vivo debromination, because these BDE 

congeners were absent in both the fish food and in the tissues from the control group. It was 

proposed that BDE 140 was directly formed from the metabolic debromination of BDE 209. 

Evidence for this is that the bromine substitution pattern of BDE 140 could not have been 

derived from the substitution pattern of the other two higher-brominated congeners tested (e.g., 

BDE 190 or BDE 183). 

 Common carp that were fed a diet spiked with either BDE 99 or BDE 183 for 62 days 

showed significant in vivo debromination of both BDE congeners (Stapleton et al., 2004). 

Six experimental and two control groups (12 fish to a group) were fed spiked or clean fish 

pellets, respectively. One fish from each tank was sampled on Days 0, 5, 10, 20, 30, 44, 62, and 

during 37 days of depuration following dietary exposure to BDEs. The gastrointestinal tissues 

and liver were dissected from the fish and analyzed for BDEs as was the remaining homogenized 

fish sample. BDE 99 was debrominated within carp tissues (gastrointestinal tract and liver) to 

BDE 47 at a rate of 9.5% ±0.8% of the BDE 99 dietary concentrations. BDE 28 was detected in 

low concentrations in carp tissues from exposure Days 20 to 62, suggesting that BDE 47 was 

further debrominated, although this was not proven. BDE 183 was debrominated to BDE 154 

and another unidentified hexaBDE congener. The rate of biotransformation to hexaBDE 

congeners was estimated to be 14.2% ±5% of the BDE 183 dietary concentrations. The site of 

debromination was the intestinal tract. The authors postulated that the metabolic debromination 

of BDE 99 and BDE 183 could be mediated by either intestinal microflora, and/or by 

endogenous enzyme systems such as the hepatic mono-oxygenase enzymes.
3-24

Stapleton et al. (2006a) provided additional confirmational evidence of metabolic 

debromination of BDE 209 in fish. Sixty juvenile rainbow trout were randomly placed in 

aquaria holding 15 fish each (with one aquarium containing a control group). The exposed fish 

were fed fishmeal spiked with 939 ±14 ng BDE 209/g wwt for a period of 5 months, which is 

equivalent to a dose of 9.4 μg BDE 209/kg body weight/day. After 5 months of exposure, the 

fish were sacrificed, and fresh blood was drawn from the dorsal aorta. Livers were removed and 

analyzed for PBDE congeners and assayed for microsomal activity. Approximately 401 ±68 ng 

BDE 209/g wwt PBDEs accumulated in the fish liver during the experiment, indicating that the 

liver was the primary accumulating tissue. In the pooled serum, collected prior to the 

experiment, it was determined that BDE 209 averaged <2.4 ng/g serum, whereas during the 

experiment, the concentration in pooled blood ranged from 26−40 ng/g serum. Steady-state 

serum concentrations were achieved in the last 2 months of the experiment. Debromination was 

observed with the appearance of lower brominated congeners that increased in concentration in 

the fish tissues throughout the exposure period. Three nonaBDE, six octaBDE, and four 

heptaBDE congeners were present in the exposed fish tissues that were not present in the spiked 

food mixture. High resolution GC/MS confirmed the identification of the following BDE 

congeners in the fish tissue: BDE 188, BDE 201, BDE 202, BDE 207, and BDE 208. While 

BDE 207 and BDE 208 initially increased in tissue concentration, concentrations decreased over 

the last 2 months of the exposure period. The authors speculated that the BDE 207 and BDE 208 

continued to be metabolized, albeit, the study did not identify specific metabolites. A mass 

balance by Stapleton et al. (2006a) indicated that debromination of BDE 209 to less brominated 

congeners occurred in fish liver and the metabolites were subsequently transported throughout 

the body by means of the circulatory system.

 In summary, four major fish feeding studies have demonstrated the potential for in vivo 

debromination of PBDEs. The higher brominated BDE 209 and the decaBDE commercial 

formulation have been successfully debrominated through a metabolic pathway in fish. The 

exact mechanism remains unknown. Current theory suggests that enzymes involved in thyroid 

hormone catabolism, as well as hepatic enzymes, may play a role in the biotransformation of the 

higher brominated species to lower brominated congeners. It is not clear whether all fish 

species, both marine and freshwater, have the innate propensity for biotransforming PBDE, but 

debromination of PBDEs has been observed in juvenile rainbow trout, juvenile lake trout, and 
3-25

common carp (all of which are freshwater species). It seems that bromine atoms in the meta

positions, positions 3 and 3’, on the BDE molecule are most easily removed. La Guardia et al. 

(2007) provide evidence that debromination in fish and crustaceans occurs in the natural 

environment outside of laboratory conditions. La Guardia et al. (2007) compared the PBDE 

congener profiles in the sediments near the outfall of a sewage treatment plant with the PBDE 

profiles found in the tissues of sunfish (Lepomis gibbosus), creek chub (Semotilus 

atromaculatus), and crayfish (Cambarus puncticambarus). Similar BDE congener profiles were 

observed in the sewage treatment plant sludge, the surficial sediments at the outfall of the 

wastewater discharge, and downstream from the sewage treatment plant. Tri through decaBDE 

congeners were detected in the samples, with BDE 209 dominating the sediments. The resident 

sunfish tissues contained BDE 209 and also congeners not present in the sewage sludge and 

sediments. These were two octa (BDE 201, BDE 202) and three hepta (BDE 188, BDE 184, and 

BDE 179) congeners. La Guardia et al. (2007) proposed that the presence of these congeners in 

the Sunfish was suggestive of in vivo synthesis from the metabolic debromination from 

BDE 209. 

3.4.2.2. Evidence for Debromination in Rats 

In vivo debromination of PBDEs has been observed in a rat feeding study. Huwe and 

Smith (2007) found evidence of the metabolic debromination of BDE 209 in male 

Sprague-Dawley rats fed a commercial formulation of decaBDE (98.5% BDE 209). Other 

congeners detected in the formulation included nonaBDEs, octaBDEs, and a trace of BDE 183. 

DecaBDE was mixed in corn oil to a concentration of 18.9 μg of BDE 209/mL of oil. Eighteen 

rats were fed an oral, daily dose of 3.8 μg in 200 μL of oil/rat (equivalent to 0.3 μg/g-day of the 

total diet) for a period of 21 days. A control group was fed the standard rat diet over the same 

period. Experimental rats were sacrificed in groups of three on Days 0, 3, 7, 10, 14, and 21, 

following the cessation of dosing with BDE 209. Prior to sacrifice, daily samples of urine and 

feces were collected from all rats. After sacrifice, samples of the blood (plasma), the liver, the 

gastrointestinal tract, and the remaining carcass were collected from each rat and homogenized 

prior to sample analysis by an isotope dilution GC/MS method. BDE 209, nonaBDE, and 

octaBDE congeners were found to accumulate in the rat liver of the dosed animals at amounts 

that were 2−3 times higher than in other tissues. Evidence of metabolic debromination of 
3-26

BDE 209 to lower congeners was observed from an apparent 160% increase in the tissue 

concentration of BDE 197, BDE 201, and BDE 207 as compared to levels in the feed. Huwe and 

Smith (2007) postulated that the possible formation of BDE 197 and BDE 207 resulted from 

removal of bromine atoms from the meta positions on the BDE 209 molecule. The formation of 

BDE 201 was postulated to occur from the debromination along para and meta positions of the 

BDE 209 molecule.

The rat may serve as an indication of the possibility that metabolic debromination of 

BDE congeners occurs generally in mammals, including humans. However, this remains to be 

proven. 

3.4.2.3. Evidence for Debromination in Birds

There is evidence for the metabolic debromination of PBDEs in chickens (Pirard and 

Pauw, 2007) and starlings (Van den Steen et al., 2007). Pirard and Pauw (2007) fed seven

Sexaline hens a diet containing 3.4 mg/kg pentaBDE formulation (De-71, Great Lakes 

Chemicals) for 14 weeks. Egg samples and daily excreta samples were collected during the 

experiment. At the end of 14 weeks, the hens were sacrificed, and samples of fat and liver were 

taken for chemical analysis. With regard to tissue and egg distributions, 3,030, 3,711, and 

2,826 ng/g lipid-adjusted total PBDEs were detected in the liver, adipose tissue, and eggs, 

respectively. Pirard and Pauw (2007) derived BCFs for select BDE congeners by dividing the 

congener concentrations detected in abdominal chicken fat to the congener concentrations in the 

chicken feed. The estimated BCFs were as follows: BDE 47 = 0.7; BDE 100 = 1.8; 

BDE 99 = 0.6; BDE 154 = 2.2; BDE 153 = 2.0; and BDE 183 = 1.0. The authors investigated 

absorption/excretion percentages of BDE congeners excreted by the second week of dosing. 

Higher proportions of lower brominated compounds (i.e., BDE 47, BDE 100, and BDE 99), were 

found in chicken excreta compared to hexa- and heptaBDEs. Pirard and Pauw (2007) postulated 

that because it was unlikely that the high amount of BDE 47 found in excreta came from the 

fraction unabsorbed in the intestinal tract, the excess BDE 47 in chicken excreta was evidence of 

formation from the reductive metabolic debromination of congeners BDE 99 and BDE 100. The 

authors further speculated that BDE 153 could be debrominated to form BDE 99 and BDE 154 

could be debrominated to form BDE 99 and BDE 100.
3-27

Van den Steen et al. (2007) studied starlings for the bioaccumulation and tissue 

distribution of BDE 209. Four adult male starlings were housed in a large outdoor aviary and 

exposed to a solution of BDE 209 in peanut oil through a silastic tube implant. The exposed 

group received an implantation dose of 46.8 μg BDE 209/day for 76 days, and a control group 

(n = 3) received an implant filled with unfortified peanut oil over the same time frame. During 

the exposure period, 300 μL of blood was taken from each bird every 3−7 days. Following the 

exposure period, the experimental birds were euthanized, and the pectoral muscle and bird liver 

were excised for analysis of BDE 209. It was found from analyzing the silastic tubes that only 

50% of the total dose of BDE 209 diffused from the silastic tube implant into the bird. The 

half-life of BDE 209 in the blood of the starlings was estimated to be 13 days (95% confidence 

interval: 11 to 18 days). BDE 209 accumulated in muscle tissue at a rate twofold higher than in 

liver explained by the higher metabolic activity of the liver. In addition to BDE 209, other BDE 

congeners were detected in bird tissues. The detection of substantial amounts of BDE congeners 

BDE 208, BDE 207, BDE 206, BDE 197, BDE 196, and BDE 183 in liver and muscle was 

evidence of metabolic debromination of decaBDE in avian tissues (Van den Steen et al., 2007). 

3.4.2.4. Evidence for Debromination in House Cats

Dye et al. (2007) reported on the possible metabolic debromination of PBDE in house 

cats. The purpose of this study was to investigate the relationship between the incidence of 

feline hyperthyroidism in cats and the dietary ingestion of dry cat food contaminated with PBDE 

in addition to ingestion of PBDE contaminated house dust. The congeners BDE 47, BDE 99, 

BDE 207, and BDE 209 were most frequently detected in blood serum of 23 exposed cats and 

were related to their dietary intake. The cats consuming only canned wet cat food (mostly 

composed of fish) (n = 4) had relatively little BDE 207 or BDE 209 present in serum. BDE 47 

and BDE 99 were present in the highest concentrations in the serum of canned food eaters. 

Conversely, in cats consuming only dry-food (n = 8), BDE 209 dominated, and the detected 

congeners had the following distribution: BDE 209 > BDE 207 > BDE 47 > BDE 99. The 

remaining cats that consumed both food types (n = 11) exhibited a mix of BDE congeners in 

serum with no one congener dominating the others. 

The evidence for possible metabolic debromination of BDE 209 in house cats stemmed

from a comparison of the BDE congener distributions present in serum to the BDE congeners 
3-28

profiles in the dry and wet cat food. The contamination of PBDEs in dry cat food reflected the 

congener profile of decaBDE, with BDE 209 representing 83−93% of total PBDE present in the 

feed. Because of the high content of BDE 209 in dry cat food, BDE 209 dominated serum in 

cats only consuming the dry food. BDE 209 accounting for 4.2%, 21%, and 30% of serum 

PBDE levels in house cats consuming canned, mixed, and dry food, respectively. It was noted 

that BDE 207 was consistently present in serum in significant concentrations of the dry food 

eaters as compared to the consumers of the other food types. BDE 207 accounted for 4.5%, 

9.8%, and 17% of the PBDE levels detected in cats consuming canned, mixed, and dry 

food-eaters, respectively. BDE 207 was present in the dry food eaters at approximately 50% of 

the total concentration of BDE 209, which is uncharacteristic of the decaBDE congener profile 

(BDE 207 is approximately 1% of BDE congeners present in decaBDE), and was not the pattern 

observed in the wet food eaters. Moreover, the ratio of BDE 207 to BDE 209 in cat serum was 

relatively constant in all dry cat food eaters. The authors regarded these data as possible 

evidence for the metabolic debromination of BDE 209 to form BDE 207. These data are only 

suggestive of the in vivo debromination in house cats, but if confirmed through replicate studies, 

would generally imply the possibility for metabolic debromination of BDEs in humans. 

3.4.3. Abiotic Degradation of PBDEs

Abiotic degradation predominantly occurs in the atmosphere and on soil surfaces. The 

energy of sunlight can degrade PBDEs in air and soils via photolysis, and the presence of the 

hydroxyl radical in air can deplete some PBDEs present in the atmosphere. The following is a 

brief review of these processes.

3.4.3.1. Photodegradation of PBDEs

Several studies have shown that higher brominated BDE congeners can photodegrade to 

form lower brominated congeners as photochemical byproducts. The photodegradation is 

defined as the photochemical transformation of a molecule into lower molecular weight 

fragments, usually in an oxidation process (IUPAC, 1996). This term is widely used in the 

destruction (oxidation) of pollutants by ultraviolet (UV)-based processes; e.g., the absorption of 

photons present in wavelengths found in sunlight, i.e., UV radiation. Photodegradation of PBDE 
3-29

occurs from the removal of a bromine atom on the PBDE molecule, thus transforming the higher 

molecular weight BDE to lower brominated congeners. 

Fang et al. (2008) reported on the experimental photodegradation of BDE 28, BDE 47, 

BDE 99, BDE 100, BDE 153, and BDE 183 dissolved in hexane. Known concentrations of the 

BDE congeners (approximately 150 ng/mL) present in solvent were exposed to >290-nm UV 

wavelengths in a photo reactor consisting of a 500-W mercury lamp, quartz tubes, Pyrex glass 

tubes, and a water cooling system.. The experiments were performed in triplicate, and a control 

was used that consisted of BDE congener solutions placed in total darkness. Photodegradation 

occurred for all studied congeners, and generally followed pseudo-first-order kinetics. The 

photodegradation half-lives of the BDE congeners discerned from this study were 

BDE 28 = 4.97 hours; BDE 47 = 2.53 hours; BDE 99 = 0.32 hours; BDE 100 = 6.46 hours; 

BDE 153 = 0.29 hours; and BDE 183 = 0.26 hours. Generally, the higher brominated BDE 

congeners degraded at a faster rate than the lower brominated congeners. It was concluded that 

the main decomposition mechanism induced by photolysis was reductive debromination to form 

lower brominated BDE congeners. Fang et al. (2008) observed the following photodegradation 

products: 

 

• BDE 28 (triBDE) photodegraded to form BDE 8 and BDE 15 (both diBDEs); the diBDEs 

further degraded to form the monoBDE congeners BDE 1 and BDE 3 

• BDE 47 (tetraBDE) photodegraded to form two tribrominated species (BDE 17 and 

BDE 28), which in turn, further debrominated to form dibrominated BDE congeners

• BDE 4, BDE 8, and BDE 15 debrominated to form monoBDE congeners BDE 1 and 

BDE 3 

• BDE 99 (pentaBDE) photodegraded to form the tetraBDE congeners BDE 66, BDE 49, 

and BDE 47 (tetraBDE), which subsequently underwent further photolysis 

• BDE 100 (pentaBDE) photodegraded to form BDE 75 and BDE 47 (both tetraBDEs) 

• BDE 153 (hexaBDE) photodegraded to form BDE 99, BDE 101, and BDE 117 

(pentaBDEs)

 

Kajiwara et al. (2007) reported on experimentally photodegrading BDE 209 contained in 

flame retarded plastic. The purpose of their study was to investigate the potential for the 

photolytic debromination of the brominated flame retardants (BFRs) decaBDE and 
3-30

decabromodiphenyl ethane (DBDPE) in treated plastics. Four different flame retarded plastic 

samples were evaluated, including pure high impact polystyrene (HIPS) not treated with BFR, 

HIPS treated with decaBDE (0.15% by weight, HIPS treated with DBDPE (about 0.10% by 

weight), and used TV casing made from HIPS. It was noted that DBDPE is a chemical substitute 

for decaBDE and has a chemical structure similar to BDE 209. The plastic samples were 

dissolved in toluene and shaken overnight to be completely mixed. The toluene was evaporated 

by air drying in a dark room. After drying, solid materials, including TV casing, were pulverized 

in a liquid nitrogen chamber. The pulverized plastic samples were initially passed through a 

300 μm sieve and then a 10 μm sieve. The plastic powder collected between two sieves 

(106−300 μm) was used in the UV irradiation experiment. Aliquots of the plastic powder (0.3 g) 

were placed in sealed quartz tubes and subsequently exposed to natural sunlight at time intervals 

of 0, 7, 14, 28, 56, and 112 days. Controls (tubes of the same powdered plastic samples) were 

placed in a temperature-controlled dark room. Following each test period, samples were 

extracted and then analyzed using high resolution gas chromatography (HRGC) coupled with 

high resolution mass spectrometry (HRMS) in the selective ion monitoring mode. Photolysis of 

BDE 209 in the decaBDE-treated HIPS samples was observed after 1 week of exposure to 

sunlight. The BDE 209 concentration in the plastic decreased, and hepta-, octa-, and nonaBDE 

congeners were formed as products of photolysis. The mechanism of the photodegradation of 

BDE 209 was postulated to be by the process of eliminating bromine atoms on the molecule. 

Approximately 80% of the BDE 209 had degraded after 112 days of exposure to sunlight, but 

only 5% of the BDE 209 mass was converted to lower brominated BDE congeners, suggesting 

the formation of unknown products of photolysis. Tri- through octa-polybrominated 

dibenzofurans (PBDFs) appeared in the irradiated samples, suggesting that the photolysis of 

BDE 209 formed PBDFs, but not in sufficient quantities to close the mass balance of the 

products formed. No clear pattern of photodegradation was observed with the other plastic

samples. No loss of BDE 209 was observed in the control samples. This study demonstrated the 

photodegradation of BDE 209 in treated HIPS by natural sunlight. 

Stapleton and Dodder (2008) experimentally photodegraded decaBDE present in house 

dust by exposure to natural sunlight. Two different house dust materials were obtained from the 

National Institute of Standards and Technology in Maryland. The first material was typical 

indoor house dust known as Standard Reference Material (SRM) 2585. During chemical 
3-31

analysis, it was found to be contaminated with a variety of BDE congeners, ranging from tri- to 

decabromo-substituted congeners. The second dust sample was identified as SRM 2583 and was 

a standard reference material used in the analysis of metals and organics in dust. The SRM 2583 

dust sample was precleaned to be free of PBDE contamination and then spiked with 1.298 g of a 

solution of BDE 209 dissolved in toluene. The sample was air dried to evaporate the toluene, 

resulting in a concentration of 2,180 ng BDE 209/g dwt in the SRM 2583 dust sample. The 

SRM 2585 dust sample was not precleaned nor spiked with BDE 209. Aliquots of both SRM 

dust samples were placed into two 4.5 mL methylacrylate chambers (referred to as UV cuvettes), 

which are routinely used for measuring UV absorbance of test materials. The UV cuvettes

containing the dust samples were exposed to natural sunlight outdoors in a tray lined with 

aluminum foil. Except during periods of precipitation, the cuvettes were placed outside daily, 

Monday through Friday, from approximately 9:00 am to 4:00 pm until 200-hour cumulative 

exposure to sunlight was achieved. Samples were transferred to a laboratory where they were 

extracted and analyzed for PBDE using high resolution gas chromatography coupled with high 

resolution mass spectrometry operated in the electron capture chemical ionization mode. The 

average intensity of sunlight over the duration of the experiment was 545 watts per square meter

(W/m2

), with a range of 61 to 929 W/m2. The average outdoor temperature was 26.9°C and 

ranged from 18.7 to 32°C. Photodegradation of BDE 209 was observed in both SRM dust 

samples after 100-hour exposure to natural sunlight. The first-order BDE 209 photodegradation 

rates were calculated as 2.3 × 10−3 and 1.7 × 10−3 per hour in the spiked (SRM 2583) and natural 

(SRM 2585) dust samples, respectively. The octabrominated congeners, BDE 201 and 

BDE 202, were observed in both reference dust samples to be products formed from the 

debromination of BDE 209 induced by photodegradation. In the SRM 2583 (spiked) sample, 

additional congeners were formed BDE 183, BDE 197, BDE 202, BDE 203, BDE 206, 

BDE 207, and BDE 208. From these data, the authors calculated a mass balance based on the 

photodegradation of BDE 209 in the spiked dust samples. After 200-hour exposure to natural 

sunlight, the initial concentration of BDE 209 was observed to have decreased by approximately 

38%. The authors concluded that 35% of the decrease in the concentration of BDE 209 was due 

to debromination and the subsequent formation of lower brominated BDE congeners, and 3% of 

the decrease in concentration was postulated to have been caused by the volatilization of 

BDE 209. 
3-32

Rayne et al. (2006) photodegraded BDE 153 in acetonitrile, distilled water, and seawater 

at ultraviolet radiation wavelengths (UV radiation) of 302 nm. BDE 153 dissolved in acetonitrile 

and irradiated for 5 minutes formed three primary photodegradation products: (1) pentaBDE 

isomers BDE 99, BDE 101, and BDE 118 (20% yield); (2) brominated dibenzofuran congener 

1,2,4,7,8-PeBDF (30% yield); and (3) three nonspeciated tetrabrominated 2-hydroxybiphenyls 

(20% yield). Continued irradiation up to 1 hour caused the photodegradation of the pentaBDE 

congeners to form the tetraBDE isomers BDE 47, BDE 49, BDE 66, and BDE 77. The 

irradiation of BDE 153 in distilled and/or seawater produced the same tetraBDE isomer 

byproducts, albeit, less efficiently than acetonitrile. Rayne et al. (2006) were unable to conclude 

firmly whether or not the photodegradation of PBDEs in aquatic systems under natural 

conditions is a viable process. 

Bezares-Cruz et al. (2004) decomposed BDE 209 to form lower brominated BDE 

congeners through natural sunlight at wavelengths of 300, 305.5, 311.4, 317.6, 325.4, 332.4, and 

368 nm. DecaBDE formulation (approximately 97% BDE 209) was dissolved in hexane to 

create three solutions of BDE 209, ranging in concentrations from 6.92 × 10−6

 to 6.14 × 10−6 μM. 

Control samples were prepared in the same manner but kept in darkness. Samples were exposed 

to sunlight on clear days in the summer and fall of 2003. BDE 209 dissolved in hexane and 

exposed to sunlight photodegraded within minutes. After 30−45 minutes of exposure to 

mid-afternoon sunlight on 7/2/2003 and 10/23/2003, the BDE 209 concentration was reduced to 

approximately 5% and 1% of the initial concentration, respectively, corresponding to a 

pseudo-first-order reaction rate of 1.11 × 10−3/s and 1.86 × 10−3/s. The higher reaction rate in 

July was due to the increased intensity of solar flux as compared with October. There was no 

evidence of degradation of BDE 209 in the control solutions not exposed to sunlight. The solar 

irradiation of BDE 209 dissolved in hexane catalyzed the reductive debromination of the 

congener. Forty-three PBDE congeners of various bromine substitutions were formed during 

different times of exposures to sunlight. After 5 minutes of solar irradiation, the disappearance 

of BDE 209 was matched by the initially rapid formation of nonaBDE congeners. The other 

congeners (octaBDEs, heptaBDEs, and hexaBDEs) accumulated successively over the 60-minute 

exposure period. OctaBDEs were transformed to heptaBDEs that were then transformed to 

hexaBDEs. With respect to the possibility of the photodegradation of BDE 209 in natural 

waters, Bezares-Cruz et al. (2004) postulated that the photochemical reaction would be expected 
3-33

to be somewhat attenuated by sorption of BDE 209 onto colloidal particles in the water column, 

and by the light attenuation properties of humic materials in aquatic systems. Furthermore, the 

presence of hydrogen donors necessary to invoke the reaction would likely be at lower 

concentrations in natural waters as compared to hexane.

Söderström et al. (2004) reported on the experimental photodegradation of BDE 209 in 

various matrices, including toluene, silica gel, sand, soil, and sediment. UV-exposure 

experiments were conducted both in the laboratory with artificial UV light (all matrices) and 

under natural conditions with outdoor sunlight (sand, soil, sediment). To begin the experiment, a 

0.5 g sample of each of the matrices was placed into Pyrex tubes and fortified with a solution of 

10.5 ng/μL decaBDE dissolved in toluene. The toluene was then allowed to evaporate while the 

samples were kept in the dark. For the exposure to artificial light, the samples were placed in an 

apparatus consisting of four mercury UV lamps equipped with filters to mimic the sunlight 

spectra in the UV range of 300−400 nm. The irradiance intensity from the UV lamps was 

estimated to have been 1.6 mW/cm2. The exposure of soil and sediment was extended to an 

additional 121 and 244 hours. For the conditions of natural sunlight, the Pyrex tubes containing 

fortified samples of each matrix were placed in a tray and put on the roof of the laboratory during 

the month of July 2007. Maximum UV irradiance from the sun at mid-day was measured to 

have been 2.3 mW/cm2. Exposure durations of the samples ranged from 0 to 96 hours. All 

samples were transferred to a laboratory, extracted, and analyzed using HRGC/HRMS. 

Söderström et al. (2004) found that BDE 209 photolytically degraded from exposure to both 

artificial and natural sunlight. With exposure to artificial light, photodegradation was more rapid 

and complete when BDE 209 was associated with toluene or silica gel than with sand or 

sediment. BDE 209 in toluene or silica gel degraded to 1% of the initial BDE 209 concentration 

over 8 hours when exposed to artificial light, but degraded to about 21% and 57% over 96 hours 

when associated with sand and sediment. By comparison, natural sunlight degraded BDE 209 in 

sand and sediment to about 36% and 43% over a 96-hour exposure. The half-life of BDE 209 

varied by matrix and artificial or natural sunlight. The half-life of BDE 209 in toluene or silica 

exposed to artificial light was less than 0.25 hours. By comparison, BDE 209 in sand, sediment, 

and soil had estimated half-lives of 12 hours, 40−60 hours, and 150−200 hours, respectively, 

when exposed to artificial light. For exposure to natural sunlight, the half-lives of BDE 209 in 

sand and sediments were 37 hours and 80 hours, respectively. Söderström et al. (2004) reported 
3-34

that the photolytic debromination of BDE 209 formed lower brominated BDEs in both artificial 

and natural sunlight conditions. These included BDE 47 (silica gel only), BDE 100 (toluene 

only), BDE 119 (toluene, sand, sediments, and soil), BDE 99 (toluene and silica gel), BDE 154 

(all matrices), BDE 153 (toluene, sand, sediments and soil), BDE 140 (all matrices except soil), 

BDE 128 (all matrices), BDE 183 (all matrices), and BDE 206, BDE 207, and BDE 208 (all 

matrices).

Sánchez-Prado et al. (2005) investigated the photodegradation of pentaPBDE at two UV 

irradiation intensities. The pentaBDE technical formulation was dissolved in cyclohexane to a 

concentration of 10 μg/mL. The pentaBDE mixture was composed of the BDE congeners 

BDE 47 (4.1 μg/mL), BDE 85 (0.1 μg/mL), BDE 99 (1.2 μg/mL), BDE 100 (4.1 μg/mL), 

BDE 153 (0.23 μg/mL), and BDE 154 (0.34 μg/mL). A 5 mL aliquot of the solution was 

adsorbed to 100 μm polydimethylsiloxane fibers and placed in a laboratory photo reactor 

equipped with two low-pressure mercury lamps (8−10 W, 254 nm) for up to 1 hour. Equal 

control samples were stored in complete darkness. Chemical analyses showed no degradation of 

the BDE mixture had occurred in the controls. Debromination of the BDE congeners occurred 

over both intensities of UV irradiation. The reaction rate was observed to be dependent on the 

degree of bromination (e.g., increasing rate of photodegradation with increasing number of 

bromine atoms on the molecule). The reaction rate was observed to be independent of the 

irradiation intensity. Approximately 21 BDE congeners were observed as the degradation 

products from the irradiation of the mixture of BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, 

and BDE 154. Based on sequential formation of these products, Sánchez-Prado et al. (2005) 

postulated that the principle photodegradation pathway was reductive debromination. 

The Wilford et al. (2008) study provides indirect, but weak evidence that 

photodegradation of higher brominated BDE congeners may occur on atmospheric particles

under natural sunlight conditions. Air particulates were sampled over a 1-month period 

(April 17−May 20) in 2004 at a semirural area of the United Kingdom. The area was a 

government meteorological workstation located in grasslands 4 km from the city of Lancaster. 

PBDEs in the particle phase were collected on glass fiber filters continuously, 24-hour/day. 

BDEs in the vapor phase were collected on polyurethane foam plugs (PUFs), but only over a 

7-day period. The purpose of PUF sampling was to confirm the dominance of particle bound 

BDE 209 during the cool month of spring. Mean air sampling volumes were 380 and 360 m3 for 
3-35

particle phase only samples, and for particle and vapor phase samples, respectively. An 

additional 9-day quality control sample was taken to confirm that the intended internal and 

recovery standards (BDE 181 and BDE 190) were not present in the atmosphere. Samples were 

collected daily and stored in amber glass containers at −20°C until analysis. Samples were 

analyzed by gas chromatography coupled with an electron capture detector for the higher 

brominated compounds BDE 197, BDE 196, BDE 208, BDE 207, BDE 206, and BDE 209. 

BDE 32, BDE 17, BDE 28, BDE 35, BDE 37, BDE 75, BDE 49, BDE 71, BDE 47, BDE 66, 

BDE 77, BDE 100, BDE 119, BDE 99, BDE 85, BDE 154, BDE 153, BDE 138, BDE 166, 

BDE 181, BDE 183, and BDE 190 were analyzed using gas chromatography coupled with mass 

spectrometry in negative ion chemical ionization mode. From the particle phase of sampling, 

mean air concentration of BDE 209 was found in a range from ND to 100 pg m−3. The mean of 

total PBDEs ranged from ND to 400 pg m−3

. The investigators had expected BDE 209 to 

dominate all higher brominated congeners present in the particle phase. However, in this study, 

the nonaBDE congeners ([BDE 207] > [BDE 208] ≥ [BDE 206]) were found to be at much 

higher concentrations than BDE 209. Wilford et al. (2008) suggested a possible explanation for 

nonaBDE congeners dominating over BDE 209. It is possible that debromination or degradation 

of BDE 209 (i.e., actual or relative enrichment of nonaBDEs) is occurring in the atmospheric 

environment (e.g., photodegradation). 

Raff and Hites (2007) reported on the relative importance of photolysis in the removal of 

PBDEs from the atmosphere by comparing photolytic degradation with the atmospheric removal 

processes of wet and dry deposition. The platform for this comparison was the study of the 

temporal concentration profile of PBDEs in the sediments of Siskiwit Lake on an island in Lake 

Superior. Through the temporal study of Siskiwit Lake, the investigators concluded that the 

PBDEs detected in the lake sediments were entirely the result of atmospheric deposition. This 

conclusion was reinforced by the fact that there were no significant anthropogenic sources within 

Siskiwit Lake’s drainage basin (Raff and Hites, 2007). The mean concentration of BDE 209 in 

the sediment was 2,600 pg/g. This amount accounted for approximately 95% of the total mass of 

PBDEs in the most recent lake sediment layer, which corresponds roughly to the time frame of 

1995−2005. The remaining BDE congeners were BDE 47, BDE 99, BDE 100, BDE 153, and 

BDE 154 at mean concentrations of 57, 69, 20, 9, and 9 pg/g, respectively. The investigators 

examined the importance of photolysis in explaining the BDE congener profile in the lake 
3-36

sediments by comparing photolytic half-lives of BDE congeners to the first-order removal rate

constants for oxidant reactions and dry and wet deposition. Photolytic degradation rates of 

25 BDE congeners were estimated from in-house experiments in a conventional photo-reactor

exposing PBDEs to UV radiation in wavelengths of 280−350 nm. Deposition fluxes of the BDE 

congeners were calculated from the sediment concentrations and the sedimentation rate of the 

lake. The phase distribution of BDE in ambient air, along with the measured size distribution 

characteristics of atmospheric particles over Lake Superior, were used in the computation of 

atmospheric deposition. Raff and Hites (2007) assumed a mean dry particle deposition velocity 

of 0.2 cm/s. Wet deposition was calculated from estimates of atmospheric washout ratios of 

BDE congeners and the frequency and intensity of annual rainfall events. Dry vapor deposition 

of BDE congeners was estimated from their respective Henry’s law constants (refer to Raff and 

Hites [2007] for the calculation method). The atmospheric depletion rate constants of the vapor 

phase BDE congeners (from reaction with the atmospheric hydroxyl radical) were calculated 

from published experiments and application of structure-activity relation techniques, and 

assuming an average hydroxyl radical concentration of 9.7 × 105 molecules/cm3

. It was found 

that the dominance of BDE 209 in the lake sediments is explained by the atmospheric removal 

process of the wet deposition of the particle bound BDE 209. The first-order atmospheric 

removal rate constant (k) for the wet deposition of BDE 209 was calculated to be 13 × 10−6/s, 

corresponding to an atmospheric lifetime of 0.89 days. By comparison, the atmospheric removal 

rate constants for dry deposition, photolysis, and depletion by the hydroxyl radical were 

0.80 × 10−6/s, 0.047 × 10−6/s, and approximately 0, respectively, corresponding to atmospheric 

lifetimes of 14 days and >200 days for dry deposition and photolysis. Raff and Hites (2007) 

determined that photolysis is a minor atmospheric removal process for BDE 209, and almost all 

of the BDE 209 in the sediments of Siskiwit Lake is due to the atmospheric wet and dry 

deposition (dominated by wet deposition) of particle bound BDE 209. On the other hand, Raff 

and Hites (2007) determined that photolysis is the major atmospheric removal pathway for 

BDE 47 and BDE 99, having estimated atmospheric removal rate constants of 26 × 10−6

/s and 

43 × 10−6/s, corresponding to atmospheric lifetimes of 0.45 and 0.27 days, respectively. 
3-37

3.4.3.2. Reaction with the Hydroxyl Radical

Tropospheric reactions with the hydroxyl radical have been observed to be an important 

atmospheric degradation pathway for halogenated hydrocarbons. The hydroxyl radical is 

generally formed from the photolysis of ozone (O3) to form a single oxygen atom that 

subsequently combines with hydrogen. No direct measurements of atmospheric degradation 

rates of PBDEs from interaction with the hydroxyl radical could be located in the literature. 

However, atmospheric half-lives have been estimated from quantitative structure-activity 

relationships. Assuming an hydroxyl radical concentration of 5 × 105 hydroxyl radicals per cm3, 

the atmospheric half-lives of penta-, octa-, and deca-bromodiphenyl ether homologues have been 

estimated to be 29, 140, and 476 days, respectively (ATSDR, 2004). A degradation rate constant 

for BDE 99 has been estimated to be 1.27 × 10−12 cm3

/molecule/s using the same assumption of 

hydroxyl radical concentration (EU, 2001). The ATSDR concluded that atmospheric 

degradation from reaction with the hydroxyl radical is likely to be an insignificant atmospheric 

loss mechanism (ATSDR, 2004). 

 

3.5. THERMAL DECOMPOSITION OF PBDE

PBDEs are added to plastics, textiles, and other materials to inhibit combustibility and 

delay the spread of fire (see Chapter 2 for more detail). Basically, the PBDE additives interfere 

with the combustion process by forming bromine gas, which, in turn, displaces the O2 necessary 

to sustain the oxidation reactions. In addition, bromine radicals are formed that interfere with the 

chain reactions of hydrogen and hydroxyl radicals in the fire. The duality of these processes 

severely retards and constrains thermal oxidation.

The combustion of PBDE treated materials can lead to the formation and emission of 

PBDFs in the smoke (Weber and Kuch, 2003). It is believed that the molecular structure is 

suitable for PBDEs to be precursor compounds to PBDFs formation. Debromination of 

BDE 209 occurs at temperatures of about 500°C, leading to the formation of brominated 

biphenyls. Further combustion of brominated biphenyls causes PBDFs to be formed within the 

combustion gases (Webber and Kuch, 2003). Rupp and Metzger (2005) formed both PBDFs and 

polybrominated dibenzo-p-dioxins (PBDDs) as byproducts from the thermolysis of BDE 47 and 

BDE 153 in quartz ampoules heated to temperatures varying from 250−500°C for 5−10 minutes. 

OctaBDF (octabrominated dibenzofuran) and heptaBDF congeners can be formed and emitted 
3-38

during the extrusion of HIPS plastic treated with decabromodiphenyl ether/antimony 

(antimony [III] oxide at 275°C [Luljk et al., 1992]). The yields of PBDFs showed a significant 

increase as a function of the number of extrusion cycles. Lower-brominated diphenyl ethers 

were present in the emissions from the extrusion process confirming that the debromination of 

decaBDE and subsequent exchange of bromine atoms with hydrogen is a necessary step to the 

thermolytic formation of PBDFs. In a series of experiments of combusting deca-, penta-, and 

octaBDEs in quartz ampoules at 510−630°C, Buser (1987) formed PBDFs and PBDDs from the 

intramolecular cyclization reactions involving the attack of an oxygen atom on the diphenyl ether 

molecule. The initial steps to these reactions were the debromination of the parent compound 

followed by hydrogen substitution. Polybrominated benzenes and polybrominated phenols were 

also formed as combustion byproducts. 

3.6. PATTERNS OF ENVIRONMENTAL FATE OF PBDE

Commercial octa and pentaBDE consist of mixtures of BDE congeners. PentaBDE 

predominately contains BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154, and commercial 

octaBDE mainly contains hepta-, octa-, and hexaBDE congeners. Commercial decaBDEs are 

almost entirely BDE 209 with a small amount of nonaBDE. The BDE congeners are dispersed 

into the open environment whenever the commercial formulations are released from their 

manufacture, use, and disposal. The environmental fate of PBDE congeners is dictated by their 

physical and chemical properties and their propensity for biotic and abiotic transformation. 

These characteristics were reviewed in previous sections to this chapter (see Section 3.2). 

Chapter 2 reviewed the use and life cycle of PBDE formulations and presented estimates of 

environmental releases. The purpose of this section is to describe, in general terms, the fate of 

PBDEs once they are released into the environment.

PBDEs are mainly discharged into the air from production, use, and recycling of 

PBDE-treated plastics, electronics, computers, textiles, and polyurethane foam. PBDEs are 

discharged into surface waters from industrial activity and STPs. The land disposal of sewage 

sludge and industrial sludges also contributes to environmental loadings. 

Dispersion of BDE congeners in the environment is governed by their respective physical 

and chemical properties. Air and water are primary transport media, while soils and sediments 

are environmental sinks. Transport of BDEs occurs over relatively long distances, >1,000 km. 
3-39

Evidence for this comes from the presence of PBDEs in sediments and the tissues of deep 

ocean-dwelling whales and other marine mammals. In general, BDE congeners are relatively 

hydrophobic and lipophilic compounds that have low water solubility and low vapor pressures. 

BDE congeners bioaccumulate into terrestrial and aquatic food webs. The body burdens of BDE 

congeners in a wide variety of avian species, fish, insects, and aquatic and terrestrial mammals 

indigenous to geographical areas ranging from the equator to the poles also substantiate BDE’s 

propensity for long-range transport. The following sections discuss the fate of PBDEs in air, 

water, and soil. 

3.6.1. Fate of PBDEs in Outdoor and Indoor Air 

3.6.1.1. Fate of PBDEs in outdoor air 

The air compartment is the primary medium of dispersing PBDEs over large 

geographical areas. PBDE behavior in air is that of other semivolatile compounds (i.e., BDE 

congeners partition between the vapor phase and particle phase in accordance with their 

respective vapor pressures [see Table 3-6]). At standard temperature and pressure, the mono 

through tri brominated compounds are primarily in vapor phase; the tetra and penta congeners 

split between vapor and particle phases but predominate in the vapor phase. The hexa- and 

octa-brominated BDEs are mostly attached to atmospheric particles; and BDE 209 is exclusively 

adsorbed to atmospheric particles. Experimental evidence suggests that the lower-brominated

BDEs, those that predominate in the vapor phase, can be degraded in the atmosphere through a 

reaction with the hydroxyl radical, as well as through photolytic chemistry. Wania and Dugani 

(2003) investigated the propensity for long-range atmospheric transport through the atmosphere 

with the application of four long-range atmospheric transport models. Wania and Dugani (2003) 

concluded that the tetra- and pentaBDE congeners have a greater tendency to be transported long 

distances than lower and higher-brominated species. The characteristic travel distances of 

different PBDEs were calculated from the air models. The characteristic travel distance is

defined as the horizontal distance traveled by a parcel of air whereby 63% of the initial air 

concentration of PBDE is depleted (Wania and Dugani, 2003). The models showed that the 

characteristic travel distance ranged from 483−1,113 km for tetraBDE; 608−1,349 km for 

pentaBDE; and 480−735 km for decaBDE (Wania and Dugani, 2003). In theory, 
3-40

• Vapor phase mono, di, and tri brominated compounds have the highest tendency for 

long-range transport in the atmosphere because they are unsusceptible to particle surface 

deposition. 

• The tetra and penta brominated congeners more equally partition between vapor and 

particle phases and have characteristic travel distances that are somewhat less than the 

vapor phase compounds. 

• This bi-phased distribution in air increases the atmospheric removal of the moderately

brominated compounds because of the partial adsorption to atmospheric particles and 

subsequent wet and dry deposition to the surface. 

BDE 209 is almost entirely associated with atmospheric particles and, therefore, its long-range 

transport potential is theoretically diminished by the deposition processes. However, adsorption 

of BDE 209 to atmospheric aerosol particles leaves the possibility that a small fraction of the 

BDE 209 contaminated aerosol can be transported over long distances, which is dependent only 

on the fate of the aerosol (Wania and Dugani, 2003; Gouin et al., 2006). There exist significant 

gaps in the understanding of how aerosols are dispersed regionally and globally. This hinders 

the possibility of making definitive conclusions on the long-range transport potential of 

BDE 209. Aerosol particles may mimic the behavior of gases, making them less prone to surface 

deposition; this may explain the measurement of low air concentrations of BDE 209 in the Arctic 

whenever the prevailing winds emanate from urban areas in Alaska and Canada (Wang et al., 

2005). Others have observed the routine presence of the lighter, more volatile PBDE congeners 

(i.e., BDE 28, BDE 47, and BDE 99) in air over remote oceanic regions; BDE 209 has only been 

occasionally detected (Wurl et al., 2006a). It has been noted (see Section 3.4.3.1) that BDE 209 

is labile to photodegradation by natural sunlight. Photodegradation results in the loss of bromine 

atoms on the BDE 209 molecule to form lower-brominated BDE congeners as products of 

photolysis, and this may further explain the low potential for long range transport of BDE 209. 

Photodegradation is likely a major atmospheric process for degrading and transforming 

higher-brominated PBDEs to less brominated BDE congeners, and experiments have clearly 

shown that photodegration of PBDEs is induced by the spectra of UV radiation under natural 

conditions. 

Generally, BDE 47, BDE 99, and BDE 209 dominate the congener pattern when detected 

in urban air. The densely settled urban areas are geographically dispersed sources of the BDE air 

burden, and a strong negative atmospheric concentration gradient from the urban center out to 
3-41

rural and remote areas is often observed in air monitoring studies (Gouin et al., 2005). This 

implies that urban areas contribute to background air concentrations of PBDE in rural and remote 

settings and should be considered as major area-wide sources of PBDEs present in ambient air. 

Section 4.5 discusses concentrations of individual congeners and total BDEs in indoor and 

outdoor air in the United States and abroad. 

In addition to the physical processes of wet and dry surface deposition, and 

photochemical reactions and degradation/debromination, the leafy surfaces of deciduous forest 

canopies can significantly deplete PBDEs in the atmosphere via air-to-leaf transfer (Su et al., 

2007). The uptake into leaves transfers the BDE congeners from the air to the terrestrial 

ecosystem when the leaves drop to the surface. This process is seen as being both vapor phase 

and particle bound interception by the leaf during the growing season―with the vapor phase 

more easily absorbed. The more volatile, low molecular weight PBDE compounds are deposited 

primarily through gaseous diffusion and achieve a partitioning equilibrium between the air and 

the leaf. The less volatile, heavier PBDE compounds are scavenged by the leaf via dry particle 

deposition caused by particle impaction and diffusion (Horstmann and McLachlan, 1998). These 

processes of vegetative atmospheric depletion of semivolatile compounds are also observed with 

coniferous forests, grasses, and green leafy crops (Horstmann and McLachlan, 1998). Therefore, 

dense forests, open grasslands, and productive farmlands functionally reduce the air 

concentration of BDEs as contaminated air masses move over these areas. The octanol air 

partition coefficient (log Koa) of BDE congeners (see Table 3-5) is an important indicator of this 

propensity for uptake into leafy biomass with all congeners having a log Koa > 6. The 

atmospheric scavenging by forest canopies, grasses, and other leafy biomass can partially explain 

the seasonal variability of atmospheric BDEs. Long-term monitoring studies indicate higher 

concentrations of BDEs in air during the cooler winter months, and lower concentrations over 

the spring and summer (Gouin et al., 2005). An early springtime “bud-burst” effect of reducing 

the air concentrations of BDEs at the onset of spring has been documented to occur (Gouin et al., 

2005). In the spring and summer, greater diurnal variability in air concentrations of BDEs is 

observed during the warmer months, and, conversely, stable air concentrations of BDEs are 

observed during the cooler winter months. The relative stability of BDEs over rural and remote 

areas during the winter can be explained by diffusion of BDEs from urban centers induced by a 
3-42

negative concentration gradient, whereas the variability in the spring and summer reflects 

vegetative influences on the atmospheric concentrations (Gouin et al., 2005).

Large bodies of water also have the capacity to influence the atmospheric concentrations 

of BDEs by means of the air-water exchange of the contaminant. The rate of exchange of BDEs 

is dependent on the state of equilibrium between air and water concentrations. If the ratio of the 

BDE fugacity in water is greater than the BDE fugacity in air (in the vapor phase), then there is a 

tendency for the BDE to volatilize from the water into the air. Likewise, if the BDE vapor phase 

fugacity in air is greater than the dissolved phase in water, then there is a tendency for the 

transfer of BDE from air into water. Equilibrium is achieved when the air-water fugacity ratio 

≈1, in which case, there is no air-water exchange of BDEs. Phytoplankton biomass on the 

surface of oceans plays a role in influencing the fugacity gradient (Jaward et al., 2004). The 

BDEs in the dissolved phase in the water column will be drawn to partition to the carbonaceous 

biomass due to their high Kow (see Table 3-2). The dissolved phase concentration is decreased, 

which in turn, may induce diffusion from air to force the air-to-water exchange. In this case, the 

capacity of an ocean to absorb BDEs from the atmosphere is indirectly controlled by the density 

of surface biomass. BDEs have been measured in the dissolved phase and in suspended 

particulates to seawater (Wurl et al., 2006b). In coastal waters off China, concentrations of total 

PBDE in the dissolved phase ranged between 40.2 and 228.2 pg/L, and between 8.1 and 

69.1 pg/L in the suspended particulate matter (Wurl et al., 2006b). In the San Francisco estuary, 

BDE 47, BDE 99, and BDE 209 were found to be the most abundant congeners detected in the 

dissolved phase (Oros et al., 2005), whereas Wurl et al. (2006b) found BDE 28, BDE 47, and 

BDE 100 dominated total PBDEs in the open ocean with BDE 209 detected only at trace levels. 

In the freshwater of Lake Ontario, approximately 60% of the total PBDE was composed of 

BDE 47 and BDE 99, with congeners BDE 100, BDE 153, and BDE 154 each contributing 

approximately 5 to 8% of the total (Environment Canada, 2006). Section 4.2 provides further 

discussions on sediment concentrations of PBDEs. 

3.6.1.2. Fate of PBDEs in Indoor Air

 The previous discussion has focused on the fate of PBDEs in outdoor air. PBDEs are 

also detected in indoor air, especially sorbed to aerially suspended dust inside homes and office 

buildings. Human contact with PBDE contaminated dust in indoor microenvironments is the 
3-43

primary way in which humans become exposed (see Chapter 5). This brings into question the 

fate of PBDEs in indoor air microenvironments. As was discussed in Chapter 2, many products 

in the home and office have been made flame retardant with PBDEs, e.g., the back casings of 

television sets; the casings to personal computers (PC); facsimile machines, and printers, to name 

a few. Chapter 2 has shown that PBDEs may volatilize from these products and contaminate 

indoor air through product use. For example, laboratory experiments have observed 

volatilization of BDE 47, BDE 100, BDE 99, and BDE 85 from a computer workstation 

consisting of PC casings, mouse, keyboard, and printer (Kemmlein et al., 2003). Studies have 

also linked PBDE-treated TVs and furniture to indoor dust concentrations of PBDEs (Hirai et al., 

2006; Allen et al., 2008). A study of office buildings confirmed a qualitative association 

between the presence of PBDE-treated products and subsequent levels of PBDEs in indoor air 

(Harrad et al., 2008). Indoor air with the highest PBDE levels were in rooms in office buildings 

equipped with numerous desktop PCs and numerous office furniture items containing 

pentaBDE-treated polyurethane foam. The flexible polyurethane foam in furniture padding has 

been linked to indoor dust levels of pentaBDE congeners, and the TVs and PCs have been linked 

to decaBDE in dust (Hirai et al., 2006; Allen et al., 2008). 

 Volatilization of PBDEs and congeners from products into surrounding air does occur. 

Given the vapor pressures of PBDEs, much of the amount volatilized would be expected to 

partition between vapor and particle phases, with increasing sorption to dust particles in air with 

increased degree of bromination of the molecule. Partitioning is enhanced by the organic 

fraction of dust, and this creates a homogeneous distribution of PBDEs in indoor dust particles 

(Webster et al., 2009). A second possible mechanism of transferring PBDEs and BDE congeners 

from products to indoor dust is through the weathering and abrasion of PBDE-treated materials 

(Webster et al., 2009). The abrasion causes small polymers particles that become incorporated 

as house dust. However, evolution of small abrased particles containing PBDEs from this

mechanism is inconsistent and likely results in a heterogeneous distribution (Webster et al., 

2009). 

 Dust from indoor spaces is often collected by vacuuming. Such dust would inherently 

contain some amount of PBDE contamination. Once the vacuumed dust is disposed of, either it 

likely is transferred to a landfill or it is incinerated with the other household and commercial 

waste. Some PBDE dust may directly transfer from indoor spaces to the outdoor air by 
3-44

convection through open doors and windows. The indoor air levels of PBDE contaminated dust 

are expected to be highly variable from one building or house to another. This variability is 

caused by such factors as the age and numbers of PBDE-treated materials within the indoor 

spaces, the ventilation rate, the rate of fresh air exchange within the indoor spaces, and the active 

use of PBDE-treated products and materials (Hazrati and Harrad, 2006). 

3.6.2. Fate of PBDEs in Water

Once encompassed within surface waters as a result of atmospheric deposition and/or the 

direct discharge from anthropogenic source activities, the PBDEs partition between the water 

column and the sediments in proportion to their physical-chemical properties. The benthic 

sediments are a primary sink for PBDEs. PBDEs then can bioaccumulate up the aquatic food 

web beginning with benthic organisms and ending with predators at the top of the food chain, 

e.g., piscivorous fish, birds, and terrestrial mammals.

 In general, PBDE concentrations are highest in sediment samples collected downstream 

of the following: industrial/urban areas, outfalls to sewage treatment plants, and urban locations 

without heavy industries. The lowest PBDE concentrations are generally found in sediments 

collected at remote and agricultural areas. BDE 209 appears to dominate the congener profile of 

aquatic sediments. Section 4.2 includes a complete description of sediment studies that bear out 

these trends.

 Only two studies could be located that evaluated the phase distribution of BDE congeners 

in water. In Lake Michigan, Streets et al. (2006) found dissolved phase BDE congeners BDE 47, 

BDE 99, BDE 100, and BDE 66 in concentrations ranging from 0.13 to 10 pg/L (for individual 

congeners). Three congeners were detected in the particle phase: BDE 47, BDE 99, and 

BDE 100 were found at concentrations ranging from 0.18 to 1.4 pg/L. BDE 209 was not 

evaluated in this study. Wurl et al. (2006a) evaluated the phase distribution of BDE 28, BDE 47, 

BDE 99, BDE 100, BDE 153, BDE 156, BDE 183, and BDE 209 in the sea-surface microlayer 

(SML) and subsurface seawater (SSW) from locations off the coast of Hong Kong. The SML of 

the seawater was defined as a 100-μm thick boundary layer between the atmosphere and the 

ocean surface that is composed of naturally occurring organic matter and μm-sized suspended 

particles. The subsurface was defined as the layer below the SML down to a depth of 1 m. Over 

all samples, total PBDE concentrations ranged from 11.3 to 62.3 pg/L in the dissolved phase and 
3-45

from 26.2 to 32.5 pg/L in the particle phase. BDE 209 was detected at trace levels in the 

dissolved phase at all sampling locations. Only the BDE congeners BDE 28 and BDE 47 were 

detected in the particle phase of the SML. Below the SML in the SSW, BDE 28 and BDE 47 

were detected at three sampling locations. BDE 99 was the only other congener detected in the 

dissolved phase of the subsurface layer. 

Once in the sediments or in the suspended organic matter in the water column, the 

PBDEs bioaccumulate into the ecological food chains beginning with the benthic organisms and 

continuing up to the top predators. The BDE congeners have a high capacity for 

bioaccumulation and biomagnification in biota as indicated by their relatively high Kow factors 

(see Table 3-2). Contamination of fish tissue with BDE congeners exposes fish-eating fish, 

piscivorous birds, terrestrial animals, and humans to BDEs via the dietary pathway. There is 

suggestive evidence that fish are able to transform higher-brominated PBDE congeners (in vivo) 

to lower-brominated PBDE congeners through the process of metabolic debromination (see 

Section 3.4.2.1 for further discussion). It has been suggested that metabolic debromination may 

cause the formation of BDE 47 and BDE 99. Piscivorous raptors bioaccumulate PBDEs through 

their habitual consumption of contaminated fish. These birds may also have the capacity for 

metabolic debromination of higher-brominated BDE congeners to lower-brominated congeners, 

although the evidence for this is highly suggestive and stems primarily from the study of 

chickens and starlings. PBDE congeners have been detected in marine mammals such as whales, 

seals, and porpoises, indicating their exposures to PBDEs from their diet. Bioaccumulation 

within a broad range of animals is an indication of the consequences of widespread PBDE 

contamination in the aquatic and marine trophic networks (see Section 3.7.1., Bioaccumulation 

in the Aquatic Environment, below, for a more detailed discussion).

3.6.3. Fate of PBDEs in Soil

 Soil is a major sink and environmental reservoir for PBDEs. Atmospheric PBDEs 

exchange between the air and the soil compartments by means of gas and particle depositional 

processes. The atmospheric gas particle partitioning of BDE congeners and the type of biomass 

covering the soil surface has an effect on the flux of PBDEs from air to soil. There are apparent 

differences in the magnitude of soil concentrations of BDE congeners in soils overlaid by 

coniferous woodlands, deciduous woodlands, and grass. When the soils are covered by 
3-46

vegetation, especially forests, the scavenging of semivolatile organic chemicals from the 

atmosphere is enhanced, which, in turn, increases their depositional flux to the terrestrial surface 

relative to deposition to bare soil (MacLeod, 2003). In consideration of gas particle partitioning 

and ground cover, Palma et al. (2002) estimated the environmental media partitioning efficiency 

of air releases of unsubstituted diphenyl ether (DE), and BDE 47, BDE 99, and BDE 209 from 

PBDE sources. Fugacity modeling suggested that approximately 68% of DE and 98% of 

BDE 47, BDE 99, and BDE 209 of the air emission will partition into the soil compartment at 

equilibrium. There is suggestive evidence that PBDEs, especially BDE 209, in soils can be 

degraded by the microbial reductive debromination, as well as the photolytic debromination in 

soils, although the viability of these processes in soils is not presently well understood (see 

Section 3.4.3 for a detailed discussion). 

On a geographical and regional scale, the PBDE concentrations in soils usually reflect a 

gradient (high to low) from the central city out to rural areas consistent with atmospheric 

measurements (Harrad and Hunter, 2006). Therefore, urban areas are regional sources of PBDEs 

to soil via the air-to-soil exchange. Cetin and Odabasi (2007) found positive relationships 

between the BDE congener profile in soil and air (r

2

 = 0.13 to 0.79, p < 0.01), thus further 

supporting the assumption of a close link between air and soil. From this evidence, it is 

concluded that the atmosphere is a major transport media for the PBDEs detected in soils. 

Some soil studies of PBDEs show a dominance of congeners BDE 47, BDE 99, 

BDE 100, BDE 153, and BDE 154 (Hassanin et al., 2004), but other soil data show that 

BDE 209 dominates total soil PBDE concentration (Cetin and Odabasi, 2007), especially in 

urban areas. However, Hassanin et al. (2004) did not include BDE 209 as an analyte in the 

chemical analysis of soil samples. It is possible that BDE 209 would have been shown to 

dominate in this study if BDE 209 had been included. BDE 209 also has been shown to 

dominate the sediment profile. Dominance of BDE 209 reflects the change from the general 

global use of the commercial penta- and octaBDE formulations to decaBDE, as well as high 

sorption and persistence of BDE 209. Further discussions of PBDEs in surface soils are found in 

Section 4.3. 
3-47

3.6.4. PBDEs in Sewage Treatment Plant Influent, Effluent, and Sludge

STP operations are likely a significant source of PBDEs to surface water leading to local 

contamination of the freshwater and coastal marine environments. The STP receives wastewater 

from homes, businesses, and, in many cases, industries, which subject the wastewater to different 

degrees of treatment before the treated effluent is discharged into surface waters. Much of the 

STP sewage sludge generated by the treatment of wastewater is disposed of on land, which, in 

turn, can lead to water pollution through soil erosion into surface waters. 

 North (2004) sampled influent, effluent, and sewage sludge for the presence of 41 BDE 

congeners at a STP in Palo Alto, CA. The STP employs tertiary treatment methods and 

processes approximately 95 × 106 L/day of wastewater generated by residents (60%), industries 

(10%), and commercial businesses and institutions (30%). The Palo Alto STP discharges treated 

effluent into to the San Francisco Estuary; therefore, the STP is likely a source of local PBDE 

contamination to the water, sediments, and biota. Results showed that more than 90% of the 

total PBDE concentrations in the STP effluent discharged to surface waters after tertiary 

treatment were composed of BDE congeners BDE 47, BDE 99, BDE 100, BDE 153, and 

BDE 209. BDE 99 dominated the congener profile. The rank order and mean concentrations of 

BDE congeners in the STP effluent were BDE 99 (11,200 pg/L) > BDE 47 (10,467 pg/L) 

> BDE 100 (1,983 pg/L) > BDE 209 (1,730 pg/L) > BDE 153 (983 pg/L). With respect to the 

sewage sludge from the STP tertiary process, BDE 47, BDE 99, and BDE 209 represented 

approximately 85% of the total concentration of BDEs detected. BDE 209 dominated the 

profile; it represented about 35% of total BDEs. The rank order and mean concentrations of 

BDE congeners in the STP sewage sludge (μg/kg dwt) were BDE 209 (1,183) > BDE 99 (944) 

> BDE 47 (757). From a mass balance perspective, North (2004) estimated that 96% of the 

PBDEs that enter the STP are adsorbed to sludge, and 4% is deposited into surface water with 

the wastewater effluent. The high percent adsorption of the BDE congeners to the STP sludge 

can be attributed to their high log Kow, which is >5.0 for all congeners (see Section 3.2.2). It is 

unknown whether BDE 209 in sewage sludge degrades photolytically or microbially to form 

lower-brominated congeners. 

Recently, the U.S. Environmental Protection Agency (EPA) conducted an updated 

National Sewage Sludge Survey (NSSS) of PBDEs and other contaminants present in sewage 

sludges in the United States (U.S. EPA, 2009). This information is used to estimate the amount 
3-48

of PBDE contamination in the sewage sludge that is applied to land as a soil amendment and 

fertilizer. The NSSS was a statistically-based survey in which sewage sludges were randomly 

selected from 74 STPs in 35 states to represent the United States as a whole. Samples were 

collected between August 2006 and March 2007. The mean concentrations (μg/kg dwt) of 

PBDE congeners in sewage sludge were BDE 47 = 709.17; BDE 99 = 716.36; BDE 153 = 68.33; 

and BDE 209 = 2,181.23. On a national basis, the dominant congener was BDE 209. 

 Knoth et al. (2007) investigated the distribution of BDE 28, BDE 47, BDE 99, BDE 153, 

BDE 154, BDE 183, and BDE 209 in sewage sludge samples from 11 STPs in Germany. 

Thirty-nine sewage sludge samples from different stages of the wastewater treatment process 

(primary sludge, secondary excess sludge, and dewatered digested sludge) were collected from 

March 2002 to June 2003. BDE 209 dominated the PBDE distribution in all STP sludges. 

BDE 209 concentrations in sludges ranged from 97.1 to 2,217 ng/g dwt, with a mean of 429 ng/g 

dwt. The sum of BDE congeners BDE 28, BDE 47, BDE 99, BDE 153, BDE 154, BDE 183 

ranged from 12.5 to 288 ng/g dwt (mean 126 ng/g dwt). The BDE congener profile remained 

rather static from one sludge type to another. Knoth et al. (2007) speculated that this may 

provide evidence for the biotransformation of BDE 209 to lower BDE congeners. With over half 

the sewage sludge applied to land, Knoth et al. (2007) estimated that 150 kg/acre of pentaBDE 

plus octaBDE, and 350 kg/acre of decaBDE were applied to land in 2001 from the land farming 

of contaminated sewage sludge in Germany. 

Wang et al. (2007) analyzed sewage sludge samples from 31 STPs in 26 cities in China 

for the distribution of BDE congeners BDE 17, BDE 28, BDE 47, BDE 66, BDE 71, BDE 85, 

BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, and BDE 209. The concentrations 

of the sum of all congeners excluding BDE 209 ranged from 6.2 to 57 ng/g dwt, with a mean and 

median concentration of 19.6 and 16.0 ng/g dwt, respectively. As with the studies discussed 

above, BDE 209 was the dominant congener in most of the sludge samples. The percentage of 

BDE 209 concentration to total PBDE concentration in Chinese sewage sludges averaged 

55%―with a median of 69%. BDE 209 concentrations in sewage sludge ranged from nondetect 

to 1,109 ng/g dwt (mean―70.8 ng/g dwt; median―25.5 ng/g dwt). BDE 209 was not detected 

in four sludge samples. In more than 80% of the sludge samples, BDE 209 was less than 

100 ng/g dwt. Other dominant BDE congeners in the sludges were BDE 47 (mean 24% of total 

BDE concentration), BDE 99 (mean 22%), and BDE 183 (mean 13%). Wang et al. (2007) 
3-49

regressed the data and determined correlations with BDE congener pairs. A significant and 

positive correlation (r = 0.814, p < 0.001) was found between BDE 47 and BDE 99. This 

association implied significant contamination of the sludge from the pentaBDE commercial 

formulation. 

Song et al. (2006) investigated the fate, partitioning, and mass loading of BDEs 

throughout the STP process from influent to sewage sludge to effluent. Three sets of samples 

were taken from various segments of the sewage treatment process at an STP in Windsor, 

Ontario, Canada over 3 days spaced over a 6-week period between the end of March and early 

May 2004. The BDE congeners BDE 28, BDE 47, BDE 71, BDE 99, BDE 100, BDE 138, 

BDE 153, BDE 154, and BDE 183 were evaluated in the study. BDE 209 was not included in 

the list of analytes. The congeners of commercial pentaBDE formulation (BDE 47, BDE 99, 

BDE 100, BDE 153, and BDE 154) were detected in all samples and at all stages of the STP 

process. On average, approximately 83% of total BDEs detected in the STP were composed of 

BDE 47 and BDE 99. BDE 47 and BDE 99 were found to be associated with the colloidal 

suspension of particles and the dissolved phase of organic matter in the wastewater influent. In 

the STP process, the wastewater being treated has a relatively short (12 hours) hydraulic 

residence time. The authors speculated that the short residence time combined with the low 

Henry’s Law constant and large estimated aqueous degradation half-lives would mean that loss 

of both BDE 47 and BDE 99 by volatilization or degradation during treatment would be 

negligible. Further, the high octanol water partitioning coefficients of BDE 47 and BDE 99 

(log Kow > 5) drives the compounds to partition to the wastewater solids, and BDE 47 and 

BDE 99 are removed from the wastewater with the removal of solids during the treatment 

process. This, in turn, enriches the sewage treatment plant sludges with BDEs but reduces their 

loading with the effluent. The mass balance calculations of PBDEs at this 61 million L/day STP 

were as follows. Of the 10,560 mg/day total PBDEs entering the STP, approximately 

9,609.6 mg/day (91%) ends up in the STP sludges (primary settling and activated sludge), and 

950.4 mg/day (9%) are discharged into the surface water with the treated final effluent.

Table 3-8 summarizes the BDE congener distributions in influent, sludges, and final 

effluent from various surveys of sewage treatment plants in various countries. Generally, North 

America has higher total BDE concentrations in STP sludge than Europe, with relatively high 
3-50

Table 3-8. The BDE congener distributions in influent, sludges, and final 

effluent from various surveys of sewage treatment plants in various countries

PBDE congener

(Br substitution)

STP influent

(ng/L)

STP sludge

(μg/kg dwt)

STP

final effluent

(ng/L)

Location Reference

BDE 1 NA ND ND Palo Alto, CA North (2004)

BDE 2 NA ND ND Palo Alto, CA North (2004)

BDE 3 NA ND ND Palo Alto, CA North (2004)

BDE 7 NA ND 0.016 Palo Alto, CA North (2004)

BDE 8 NA ND 0.0042 Palo Alto, CA North (2004)

BDE 10 NA ND ND Palo Alto, CA North (2004)

BDE 12 NA ND ND Palo Alto, CA North (2004)

BDE 13 NA ND ND Palo Alto, CA North (2004)

BDE 15 NA 0.62 0.008 Palo Alto, CA North (2004)

BDE 17 NA 5.7 0.19 Palo Alto, CA North (2004)

BDE 25 NA 0.62 0.0099 Palo Alto, CA North (2004)

BDE 28 NA

1.3 

13 

22 

0.266

ND

Palo Alto, CA

Ontario, CAN

North (2004)

Song et al. (2006)

BDE 30 NA ND ND Palo Alto, CA North (2004)

BDE 32 NA ND ND Palo Alto, CA North (2004)

BDE 35 NA ND 0.005 Palo Alto, CA North (2004)

BDE 37 NA 0.24 0.0038 Palo Alto, CA North (2004)

BDE 47 NA

102 

NA

NA

NA

757 

1,819

2.77

7.0*

709.17

10.5 

14 

NA

NA

NA

Palo Alto, CA

Ontario, CAN

Germany (11 STPs)

Sweden (22 STPs)

USA (74 STPs)

North (2004)

Song et al. (2006)

Knoth et al. (2007)

Oberg et al. (2002) 

U.S. EPA (2009) 

BDE 49 NA 18 0.266 Palo Alto, CA North (2004)

BDE 66 NA 21 0.217 Palo Alto, CA North (2004)

BDE 71 NA 2.8 0.043 Palo Alto, CA North (2004)

BDE 75 NA 1.0 0.018 Palo Alto, CA North (2004)

BDE 77 NA ND ND Palo Alto, CA North (2004)

BDE 85 NA

NA

34 

0.42*

0.352

NA

Palo Alto, CA

Sweden (22 STPs)

North (2004) 

Oberg et al. (2002)

BDE 99 NA

121 

NA

NA

NA

944 

2,004 

137.19

10*

716.36

11.2 

16.0

NA

NA

NA

Palo Alto, CA

Ontario, CAN

Germany (11 STPs)

Sweden (22 STPs)

USA (74 STPs)

North (2004)

Song et al. (2006)

Knoth et al. (2007)

Oberg et al. (2002) 

U.S. EPA (2009)
3-51

Table 3-8. The BDE congener distributions in influent, sludges, and final 

effluent from various surveys of sewage treatment plants in various countries

(continued)

PBDE congener

(Br substitution)

STP influent

(ng/L)

STP sludge

(μg/kg dwt)

STP

final effluent

(ng/L)

Location Reference

BDE 100 NA

19

NA

NA 

165 

289 

154.45

1.7*

2.0

2.8 

NA

NA

Palo Alto, CA

Ontario, CAN

Germany (11 STPs)

Sweden (22 STPs)

North (2004)

Song et al. (2006)

Knoth et al. (2007)

Oberg et al. (2002)

BDE 105 NA ND ND Palo Alto, CA North (2004)

BDE 116 NA ND ND Palo Alto, CA North (2004)

BDE 119 NA ND 0.014 Palo Alto, CA North (2004)

BDE 126 NA ND ND Palo Alto, CA North (2004)

BDE 138 NA

1.0 

NA

7.7 

26.1 

ND

0.096 

ND

NA

Palo Alto, CA

Ontario, CAN

Sweden (22 STPs)

North (2004)

Song et al. (2006)

Oberg et al. (2002)

BDE 140 NA 2.7 0.031 Palo Alto, CA North (2004)

BDE 153 NA

11

NA

NA 

NA

88 

193

27.31 

0.86*

68.33

0.98

1.60 

NA

NA

NA

Palo Alto, CA

Ontario, CAN

Germany (11 STPs)

Sweden (22 STPs)

USA (74 STPs)

North (2004)

Song et al. (2006)

Knoth et al. (2007)

Oberg et al. (2002) 

U.S. EPA (2009)

BDE 154 NA

7.6 

NA

NA

68 

120

18.51

0.72*

0.776 

ND

NA

NA

Palo Alto, CA

Ontario, CAN

Germany (11 STPs)

Sweden (22 STPs)

North (2004)

Song et al. (2006)

Knoth et al. (2007)

Oberg et al. (2002)

BDE 155 NA 7.1 0.073 Palo Alto, CA North (2004)

BDE 181 NA ND ND Palo Alto, CA North (2004)

BDE 183 NA

1.7

NA

10 

34

13.66

0.080 

ND

NA

Palo Alto, CA

Ontario, CAN

Germany (11 STPs)

North (2004)

Song et al. (2006) 

Knoth et al. (2007)

BDE 190 NA 1.0 0.0039 Palo Alto, CA North (2004)

BDE 206 NA 16 0.041 Palo Alto, CA North (2004)

BDE 207 NA 22 0.095 Palo Alto, CA North (2004)

BDE 208 NA 11 0.051 Palo Alto, CA North (2004)

BDE 209 NA

NA

NA

NA

1,183

363.46

11

2,181.23

1.73 

NA

NA

NA

Palo Alto, CA

Germany (11 STPs)

Sweden (22 STPs)

USA (74 STPs)

North (2004)

Knoth et al. (2007) 

Knoth et al. (2007) 

U.S. EPA (2009)
3-52

Table 3-8. The BDE congener distributions in influent, sludges, and final 

effluent from various surveys of sewage treatment plants in various countries

(continued)

PBDE congener

(Br substitution)

STP influent

(ng/L)

STP sludge

(μg/kg dwt)

STP

final effluent

(ng/L)

Location Reference

Total PBDEs

NA

265

NA

3,381

4,324

987.21

29

36 

NA

Palo Alto, CA

Ontario, CAN

Germany (11 STPs)

North (2004) 

Song et al. (2006) 

Knoth et al. (2007)

*Median concentration as opposed to mean.

NA = Not analyzed; ND = not detected.

concentrations of BDE 209. However, BDE 209 was not routinely evaluated in European sludge 

studies.

3.7. BIOACCUMLATION IN BIOTA

3.7.1. Bioaccumulation in the Aquatic Environment

The hydrophobic and lipophilic properties of BDE congeners cause aquatic organisms to 

bioaccumulate BDEs with exposures within their food web. Despite the recent prevalence in the 

use of the decaBDE commercial formulation, which principally contains BDE 209, the 

lower-brominated BDE congeners are most prevalent in the tissues of aquatic organisms. As 

discussed in previous sections, the BDE congeners enter a water body from atmospheric 

deposition, air-to-water transfer, or direct discharge from industries and sewage treatment plants. 

Once BDEs enter the aquatic system, they partition between the water column and sediments 

according to their physical-chemical properties. Moreover, within the water column, BDEs 

partition between the dissolved phase and the particle phase. Benthic invertebrates and mollusks 

living and feeding directly on contaminated sediments acquire a body burden of BDEs that is 

then passed up the food chain from bottom-feeding fish to top piscivorous fish, raptors, reptiles, 

and mammals. Mollusks internally absorb BDEs from the water column through filtering of the 

water. Fish acquire BDEs through the dietary pathway and also from the water passing through 

their gills. In this regard, trophic level bioaccumulation in marine and freshwater biota is similar 

to the pattern of bioaccumulation of other classes of lipophilic and persistent compounds such as 

the PCBs (Evenset et al., 2005; Bragigand et al., 2006; Eljarrat et al., 2004), i.e., the 
3-53

concentration of PBDEs from one trophic level to another seems to biomagnify between 

organisms. Bragigand et al. (2006) illustrated this point in the study of trophic level transfers of 

BDE congeners in an aquatic ecosystem in estuaries of France. They found that the 

bottom-feeder bivalves and worms had the lowest concentrations of BDE 47, while the water 

column feeder eels and soles had the highest concentrations. High concentrations were also 

found in flounder and mid-range concentrations in shrimp.

BDE 47, BDE 66, BDE 99, BDE 100, BDE 153, and BDE 154 have been detected in 

multiple species of marine and freshwater organisms. In most cases, BDE 209 has not been 

included in the list of analytes, and, therefore, it is often suggested that BDE 209 cannot be found 

in aquatic biota. However, BDE 209 has been detected in hardhead catfish, Atlantic stingrays, 

sharp-nosed sharks, and bull sharks off the Florida Coast (Johnson-Restrepo et al., 2005), in 

roach fish in a Baltic Sea estuary (Burreau et al., 2004), and in mysid shrimp off the Dutch coast 

(Verslyckea et al., 2005). 

For the most part, total PBDEs are highest in top predator marine mammals such as 

6,500 ng/g lipid weight (lwt) in white-beaked dolphin off the Dutch coast (Soni et al., 1998). 

However, water filtering marine mollusks appear to accumulate BDEs at the highest levels 

observed in any field survey (e.g., an average total PBDE concentration of 13,502 ng/g lwt in 

clams from the San Francisco Bay estuary [Oros et al., 2005]). Another observation that can be 

made from existing studies is that marine fish have higher overall body burdens of total PBDEs 

as compared to freshwater fish. Marine fish species tend to have total mean PBDE body burdens 

in a range of <1.0 to 1,600 ng/g lwt (Antarctic rockcod and Florida bull shark, respectively), 

whereas freshwater fish typically have body burdens from 1.0 to 300 ng/g lwt (Detroit River 

bigmouth buffalo fish and Lake Michigan lake trout, respectively) (Corsolini et al., 2006; 

Johnson-Restrepo et al., 2005; Valters et al., 2005; Streets et al., 2006). 

In most fish (ocean and freshwater) and marine mammals, BDE 47 is the major congener, 

contributing >30% to total body burden of PBDEs. The congener distribution in tissues of 

aquatic biota usually follow the order of BDE 47 > BDE 99 > BDE 100 > BDE 154 > BDE 153 

> BDE 49 > BDE 28. In the few studies where BDE 209 was measured in fish tissue (ocean 

fish), BDE 209 has ranged from <1 up to 88% of total PBDE body burden. Because of small 

sample size, these studies are insufficient to indicate the general contribution that BDE 209 

makes to PBDE in fish tissue. 
3-54

This review of PBDEs detected in fish and marine mammals illustrates the fact that 

PBDEs are transported long distances, far away from their original source. PBDEs are detected 

in deepwater marine fish and mammals and in marine ecosystems ranging from the equator to 

the poles. This provides additional evidence for the long-range transport of PBDEs. Details of 

specific studies measuring BDE congeners in freshwater, marine, farmed, and store-bought fish, 

including a tabular summary of congener-specific concentrations, are provided in Section 4.6. 

3.7.2. Bioaccumulation in the Terrestrial Environment

Field studies of the bioaccumulation of BDEs in terrestrial environment have included 

fish-eating mammals, birds of prey, and carnivorous mammals in the wild. These animals are 

top predators in the food chain and acquire a body burden of BDEs from trophic level exposures 

in a similar manner as aquatic organisms. The following provides a summary of 

bioaccumulation of PBDEs in terrestrial animals from trophic level exposures.

3.7.2.1. Bioaccumulation in Birds 

Birds of prey are ideal candidates to study the bioaccumulation of PBDEs because they 

are carnivores, typically secondary and tertiary consumers in terrestrial and aquatic food chains. 

Potter et al. (2009) reported on the distribution of PBDE congeners in 23 Peregrine Falcon 

(Falco peregrinus) eggs collected from 1993 to 2002. The eggs were from 13 nests at 

11 locations in the region of the Chesapeake Bay. Once collected, the eggs were kept frozen 

until analyzed. Results of analysis showed that the median total concentration of BDEs (sum of 

BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, BDE 196, BDE 197, BDE 206, 

BDE 207, BDE 208, and BDE 209) was 201 ng/g wwt. BDE 209 was detected in all egg 

samples at a medium concentration of 6.35 ng/g wwt. The authors noted that because of the 

short biological half-life of BDE 209 in birds (about 13 days), the presence of BDE 209 in egg 

tissues may indicate relatively recent exposures to the maternal bird (Potter et al., 2009). 

BDE 209 averaged 5.9% of the sum of the BDE congeners. In this study, the rank order by BDE 

concentration in Peregrine Falcon eggs (low to high) was BDE 153 > BDE 99 > BDE 100 

> BDE 154 > BDE 209 > BDE 183 > BDE 197 > BDE 47 > BDE 207. BDE 47 was a minor 

constituent in Peregrine Falcon eggs (on average only 4.4%). BDE 153 was the dominant 

congener present in Peregrine Falcon eggs, with a medium concentration of 26.1 ng/g wwt
3-55

(representing a mean contribution to total PBDEs of 28%). This was similar to BDE congener 

profile observed in a study of Peregrine Falcon eggs by Chen et al. (2008). The small sample 

size did not permit a time-trend analysis of PBDE concentrations in eggs. However, BDE 209 

concentrations were found to be higher in Peregrine Falcon eggs obtained from urban areas as 

compared to rural areas. For example, the authors indicated that the mean BDE 209 percent 

contribution to total PBDEs detected in Peregrine Falcon eggs collected from a rural low human 

population area was 0%, but the mean percent concentration to total PBDEs collected from a 

densely populated urban was 18.6% (Potter et al.,2009). This observation was consistent with a 

Peregrine Falcon egg study by Chen et al. (2008). 

 Chen et al. (2008) conducted a similar study of Peregrine Falcon eggs. A total of 

114 Peregrine Falcon eggs were collected for purposes of investigating the distribution of BDE 

congeners in egg tissues. Samples were collected from nests in Connecticut, Massachusetts, 

Maine, New Hampshire, Rhode Island, and Vermont between the years 1996 and 2006. Once 

collected, the eggs were transferred to solvent-rinsed glass jars and frozen until chemical 

analysis. Seventeen BDE congeners were analyzed in egg samples, including BDE 47, BDE 99, 

BDE 100, BDE 119, BDE 138, BDE 153, BDE 154, BDE 183, BDE 197, BDE 196, BDE 201, 

BDE 202, BDE 203, BDE 206, BDE 207, BDE 208, and BDE 209. Results showed that the total 

PBDEs (sum of 17 congeners) concentrations ranged from 74.5 to 6610 ng/g wwt, with a median 

of 440 ng/g wwt. PBDE congener profiles in Peregrine Falcon eggs (defined as percent 

contribution to total of 17 measured BDE congeners) were dominated by BDE 153, followed by 

BDE 99, BDE 183, BDE 209, BDE 197, BDE 207, BDE 154, BDE 100, and BDE 196; with 

lesser contributions from BDE 47, BDE 208, BDE 203, BDE 201, BDE 206, BDE 202, 

BDE 138, and BDE 119. The third most prevalent congener present in egg tissue was BDE 183. 

The authors observed that the elevated BDE-183 concentrations (a hallmark of the octaBDE 

commercial formulation) in Peregrine Falcon eggs may indicate the high biomagnification 

potential of BDE 183 in terrestrial food chains (Chen et al., 2008). BDE 47 was a minor 

contributor (<2% on average) in Peregrine Falcon eggs. Time-trend analysis suggested that total 

PBDE (sum of 17 congeners) contaminant levels in the Peregrine Falcon eggs did not change 

significantly over a 10-year period from 1996 to 2006 (Chen et al., 2008). The influences of 

rural versus urban environments on PBDE concentrations in eggs were investigated with the 

data. There was no statistically significant difference (p = 0.075) in the total PBDE 
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concentrations present in eggs from urban areas (median, 570; range, 150−1910 ng/g ww) and 

eggs from rural areas (median, 380; range, 75−3,570 ng/g ww). However, BDE 209 was 

significantly higher in Peregrine Falcon eggs collected in urban areas as compared with eggs 

collected in rural areas (p < 0.005). 

Naert et al. (2007) studied the body burden distribution of BDEs in brain and adipose 

tissue samples of Common Buzzards (Buteo buteo), Eurasian Sparrow Hawks (Accipiter nisus),

Cormorants (Phalacrocorax carbo sinensis), and Common Blackbirds (Turdus merula) from 

field samples collected at different locations in Switzerland between 2003 and 2005. Lower 

concentrations of BDE congeners were detected in the brain of avian species as compared to 

adipose tissue. The median total PBDE concentrations (sum of BDE 28, BDE 47, BDE 99, 

BDE 100, BDE 153, BDE 154, and BDE 183) in brain ranged from below the detection limit in 

Common Blackbirds to 14 ng/g wwt in sparrow hawks. The median concentrations of total 

PBDEs in adipose tissue ranged from below the detection limit in Common Blackbirds to 

709 ng/g wwt in sparrow hawks. The authors postulated that the difference in median 

concentrations between the two tissues (brain and adipose) is attributed to the blood-brain barrier 

protecting against the accumulation of BDEs in the brain. In this study, the sparrow hawk had 

the overall highest median body burden of total PBDEs of all the species: 790.2 ng/g wwt versus 

the Common Buzzard (34.55 ng/g wwt), Common Blackbird (0.82 ng/g wwt), and Cormorant 

(98.76 ng/g wwt). BDE 99 was the dominant congener detected in the sparrow hawk, accounting 

for approximately 40% of total PBDEs; BDE 47 was about 26% of total PBDEs. The sparrow 

hawk is a specialist feeder and consumes finches, sparrows, and Wood Pigeons. The Cormorant 

(having the next highest mean PBDE body burden) nests in swampy areas or near large bodies of 

water such as rivers and lakes. Its diet consists mainly of fish. BDE 47 contributed 

approximately 42% of the total body burden of PBDEs in the Cormorant. In the Common 

Buzzard, the most prevalent congeners detected were BDE 153 (29% of total), BDE 99 (23%), 

and BDE 47 (22%). The Common Buzzard preys mainly on small mammals (mice, voles, 

rabbits, squirrels, rats, and moles). The Common Blackbird, on the other hand, feeds principally 

on seeds and insects, and BDE 47 was the only congener detected in the adipose tissue of this 

species. 

Voorspoels et al. (2006a) investigated the occurrence and distribution of PBDEs in 

sparrow hawks, Common Buzzards, and owls of Belgium; the highest mean concentration of 
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total PBDEs was measured in liver of the sparrow hawk (i.e., 4,900 ng/g lwt), followed by the 

liver of the Common Buzzard (480 ng/g lwt) and owls (250 ng/g lwt). Congeners can be ordered 

according to their relative contribution to the total PBDE content as follows for Common 

Buzzards: BDE 153 > BDE 47 > BDE 99 > BDE 183 > BDE 100 > BDE 154; for Sparrow 

Hawks: BDE 99 > BDE 47 > BDE 153 > BDE 100 > BDE 183 > BDE 154; for owls BDE 153 

> BDE 99 > BDE 47 > BDE 183 > BDE 100 > BDE 154. BDE 209 was present in 6 out of 

44 liver samples and 19 out of 25 serum samples of sparrow hawks, for both Common Buzzards 

and owls combined (Voorspoels et al., 2006a). In 2 samples from sparrow hawks, BDE 209 was 

detected at a mean concentration of 17 ng/g lwt in liver, and 26 ng/g lwt in serum. In buzzards, 

BDE 209 was detected only at a frequency of 3 out of 29 (10%) in liver (mean = 79 ng/g lwt). 

However, in the serum of Common Buzzards, the frequency of detection was a surprising 16 out 

of 20 samples (80%). Chen et al. (2007a) also confirmed the presence of BDE 209 in liver of 

three Common Buzzards from Beijing, China (mean = 71 ng/g lwt). BDE 209 was the congener 

present at the highest concentration―a surprising 48% of total PBDEs. 

Later, Voorspoels et al. (2007) again evaluated sparrow hawks, Common Buzzards, and 

passerines for body burdens of BDE congeners. The authors found similar results: the highest 

levels of total PBDEs were detected in the liver of sparrow hawks, at 9,506 ng/g lwt. The liver 

of the Common Buzzard had a mean concentration of 727 ng/g lwt, which was considerably 

lower than liver of sparrow hawks. The PBDE concentrations in adipose tissues and eggs of 

passerines averaged 160 and 220 ng/g lwt, respectively. 

Because the passerine is a favorite prey of the sparrow hawk, Voorspoels et al. (2007) 

was able to observe the biomagnification potential of BDE congeners in a simple trophic system. 

The biomagnification potential (BMP) is derived as the ratio of lipid-normalized median BDE 

congener concentrations in the same tissues of both predator and prey (Voorspoels et al., 2007). 

A ratio >1 would indicate a potential for biomagnification between species. Voorspoels et al. 

(2007) calculated the mean BMP of BDE congeners when passed on through the dietary pathway 

from passerine to sparrow hawk. The rank of BMP from high to low was BDE 183 = 29, 

BDE 100 = 25, BDE 154 = 24, BDE 153 = 21, BDE 99 = 20, BDE 47 = 10, and BDE 28 = 4. 

The mean BMP for the sum of these BDE congeners was 17. It is apparent that the BMP 

increases from lower-brominated to higher-brominated BDE congeners within this simple 

terrestrial food chain. This trend generally tracks the log Kow of the BDE congeners (see 
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Table 3-2). For example, the log Kow of BDE 183 is approximately 8.3, which is considerably 

higher than the log Kow of BDE 47, which is 6.0. 

Chen et al. (2007a) reported on the distribution of BDE congeners in birds of prey (i.e., 

the Common Kestrel, the sparrow hawk, the Japanese Sparrow Hawk, the Little Owl, the Scops 

Owl, and the Common Buzzard) that were collected in the vicinity of Beijing, China. The sum 

of the mean PBDE congener concentrations was highest in the Common Kestrel. Chen et al. 

(2007a) indicated that these concentrations are among the highest levels in birds reported in the 

open literature: muscle = 12,300 ng/g lwt, liver = 12,200 ng/g lwt, and pooled 

kidney = 5,340 ng/g lwt. Kestrels dwell in the Beijing area year round, establishing nests and 

foraging in urban fringe or centers. This, in combination with their dietary habits (a preference 

for small mammals), suggests that the dietary pathway may have caused the high concentrations 

through biomagnification. Congener profiles of BDE concentrations in bird tissues indicated a 

dominance of the more highly brominated congeners. Profiles in the muscle and liver of Chinese 

Sparrow Hawks, for example, showed that BDE 153 > BDE 99 > BDE 47 > BDE 183

> BDE 154 > BDE 209 > BDE 207. This differed from the BDE profile in the tissues of 

Japanese Sparrow Hawks―BDE 99 > BDE 153 > BDE 47 > BDE 209 > BDE 207 > BDE 183. 

In the liver of Scops Owl, Long-Eared Owl, Upland Buzzard, and Common Buzzard, the 

congener profile was dominated by BDE 209 (followed by BDE 99 and BDE 153). Chen et al. 

(2007a) attributed the dominance of the higher-brominated congeners to the heavy use of 

commercial octa and deca formulations. 

She et al. (2008) recently reported on the distribution of PBDEs in piscivorous and 

omnivorous bird eggs obtained from a San Francisco Bay estuary in California, and from Gray’s 

Harbor, Washington. A total of 169 eggs were collected from the following species, combined: 

Caspian Tern (Sterna caspia); Forster’s Tern (Sterna forsterii); California Least Tern (Sterna 

antillarum brownie); and Clapper Rail (Rallus longirostris obsoletus). All tern species were 

fish-eating birds, and the Clapper Rail was omnivorous. The pattern of the five major PBDE 

congeners detected in the bird eggs (i.e., BDE 47 > BDE 99 > BDE 100 > BDE 153 > BDE 154) 

were consistent with the BDE congener patterns detected in fish. BDE 47 was the most 

abundant congener and represented approximately 60% of total PBDEs detected in bird eggs. 

BDE 209 was not detected in any egg sample. Total PBDEs in bird eggs ranged from about 

1,080 to 63,300 ng/g lwt. Median concentrations of PBDEs in Caspian Tern eggs for 2000−2003 
3-59

were 2,410, 4,730, 3,720, and 2,880 ng/g lwt, respectively, in Forster’s Tern eggs were 1,820, 

4,380, 5,460 and 3,600 ng/g lwt, respectively, and in California Least Tern eggs for 2001 and 

2002 were 5,060 and 5,170 ng/g lwt, respectively. In contrast, median PBDEs concentration in 

California Clapper Rail eggs for 2001 was 379 ng/g lwt. The rails consumed mostly invertebrate 

species. The authors concluded that the dietary consumption of PBDE-contaminated fish was 

responsible for the accumulation of PBDE congener in the eggs of fish-eating terns. 

3.7.2.2. Bioaccumulation in Terrestrial Mammals

Only a very limited number of terrestrial mammals have been studied for their body 

burden of PBDEs. Many mammals are top predators and, therefore, would be expected to 

bioaccumulate and biomagnify PBDEs from the food web. 

Voorspoels et al. (2006b) studied the red fox (Vulpes vulpes) for PBDE contamination 

and food chain bioaccumulation. The red fox is a top terrestrial predator that mainly consumes 

voles, rabbits, squirrels, and mice as prey. To the extent the prey becomes contaminated with 

PBDEs, the fox may be exposed to PBDEs that have biomagnified after moving through the 

trophic system. Voorspoels et al. (2006b) sampled tissue for the occurrence and distribution of 

PBDEs as a body burden in 33 red foxes indigenous to Belgium and then examined 

biomagnification of BDE congeners from prey to predator. The median sum of PBDE congeners 

measured in the foxes were considered low, ranging between 2.2 and 3.4 ng/g lwt in adipose 

tissue, liver, and muscle. BDE 209 was detected at a frequency of 40% in liver, 21% in muscle, 

and 15% in adipose tissue samples. In fox liver, the BDE 209 congener, on average, constituted 

70% of total PBDE concentration―a finding inconsistent with other top predator species. 

BDE 153 was the most frequently detected congener in all tissues (96−100% of all samples), 

whereas BDE 47 was detected in 33%, 40%, and 100% of adipose, liver, and muscle tissues, 

respectively. Voorspoels et al. (2006b) did not observe any evidence of metabolic 

debromination in the fox. The authors observed that total PBDE concentrations were lower in 

fox tissues as compared to voles and mice (the main diet of foxes). Voorspoels et al. (2006b) 

speculated that this may be due to the high capacity of the fox to metabolize lower-brominated 

BDEs in vivo, similar to what has been observed in the grizzly bear (Christensen et al., 2005). 

The authors postulated that this high metabolic activity might be related to the fact that no 

biomagnification of BDE congeners between prey and predator was observed in the fox 
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(Voorspoels et al., 2007). The study of PBDE distributions in foxes showed that BDE 209 does, 

however, bioaccumulate in terrestrial top predators such as the red fox (Voorspoels et al., 

2006b).

Christensen et al. (2005) reported on the influence of diet on the BDE congener 

distribution and other persistent organic contaminants in the tissues of 12 grizzly bears (Ursus 

arcos horribilis) (6 each from coastal and interior areas) in British Columbia, Canada. Dietary 

consumption of meat was estimated to be 0 to 19% and 13 to 61% of total diet for interior and 

coastal-maritime bears, respectively. Maritime bears mostly consumed salmon as the primary 

meat source, whereas the meat source of interior bears was more varied. The remaining diet of 

all bears consisted of vegetation. With the exception of PBDEs, the persistent organic chemicals

(POPs) were higher in the tissues of the bears that consumed meat. However, there was no 

significant difference in total PBDE tissue burden or congener distribution between the two 

groups of bears. Total PBDEs dominated total POP concentration in interior bears but not in 

maritime bears. The ranking by higher-to-lower total contaminant concentration of POPs in the 

tissues of interior and maritime bears showed the following pattern respectively: PBDEs > PCBs 

> HCB > HCH > CHL > DDT and PCBs > CHL > HCB > DDT > PBDEs > HCH. Both bear 

groups showed a marked difference in the mean BDE congener profile. Maritime fish-eating 

bears displayed a BDE congener pattern dominated by BDE 47 followed by BDE 209 > BDE 99 

> BDE 100 > BDE 153. In the interior bears, the tissues were dominated by BDE 209, followed 

by BDE 206 > BDE 47 > BDE 207 > BDE 208. The predominance of the lighter congener 

BDE 47 in the maritime bears suggests that PBDE exposure mainly occurs through the fish 

consumption pathway. It was postulated that the fish acquire PBDE contamination as a result of 

the long-range atmospheric transport and deposition into the marine food web. The authors 

speculated that the dominance of BDE 209 in the interior bears, whose diet was richer in 

vegetation than the maritime bears, may indicate the possibility of the PBDE exposures via the

pathway of air to plant to bear. In this paradigm, the vegetation accumulates PBDEs from local 

source air emissions of decaBDE and passes BDE 209 onto the bear through its diet. This notion 

was supported by the observations that BDE 209 is not detectable in fish, and the higher BDE 

congeners had negative bioaccumulation slopes in the tissue of bears, indicating a preferential 

exposure to local sources through their consumption of terrestrial vegetation (Christensen et al.,

2005). 
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The polar bear (Ursus maritimus) diet is chiefly composed of ringed seal. Wolkers et al. 

(2004) studied the polar cod-ringed seal-polar bear bioaccumulation pattern of PBDEs in an 

Arctic food web and found a food chain pattern consistent with what might be expected for the 

BDE 47 congener: BDE 47 concentrations in polar bear were greater than those in ringed seal, 

which in turn, were greater than those in polar cod were. From the 22 brominated compounds 

measured, only BDE 47, BDE 85, BDE 99, BDE 100, and BDE 154 were detected in polar cod. 

In addition to these congeners, the tissues of ringed seals contained BDE 66. Concentrations of 

PBDEs in ringed seals were overall higher than in polar cod. In particular, BDE 47 and BDE 99 

were approximately one order of magnitude higher in seals than in polar cod. PBDE 47 

comprised more than 90% of the total PBDEs present in ringed seals. Only BDE 47 was 

detected in polar bears. Male and female polar bears were found to have a body burden of 

BDE 47, which was 1.5 and 3 times the body burden of ringed seals, respectively. 

3.7.2.3. Bioaccumulation in Insects

Insects may acquire a body burden of PBDEs through contact with contaminated 

environmental media or through the dietary pathway. Only a very limited number of species of 

aquatic insects, caddisflies (trichoptera), and midges (diptera), have been investigated for the 

occurrence and distribution of PBDE in insect tissues (Bartrons et al., 2007). Most insect species 

undergo a number of developmental transformations from larva and pupa to adult, and they 

begin their life cycle in aquatic sediments. Insects are at the beginning of the food chain in most 

aquatic ecosystems. It has been noted that pupae emerge during periods of high fish activity and 

are an attractive source of food to many fish species (Bartrons et al., 2007). Therefore, in terms 

of movement through trophic systems and biomagnifying from lower to higher predators, insects 

can be viewed as an important beginning to contamination of PBDEs and other persistent organic 

pollutants in aquatic organisms. 

Bartrons et al. (2007) obtained 22 samples of larvae and pupae of four variants of 

caddisflies and midges from two, high-altitude mountain lakes in the Pyrenees Mountains of 

Spain. Samples of the larvae and pupae of trichoptera belonged to the polycentropodidae and 

limnephilidae families, and samples of diptera belonged to the chironomidae and 

ceratopogonidae families, thus producing samples from a total of four distinct insect types. Total 

concentrations of PBDEs in insect larva ranged from 0.65−1.68 ng/g dwt and 0−13.07 ng/g dwt
3-62

for caddisflies and midges, respectively. Total PBDEs in insect pupa ranged from 

5.17−9.32 ng/g dwt in caddisflies and 3.91−27.38 ng/g dwt in midges. In general, pupae 

contained significantly higher concentrations of total PBDE than larvae of the same taxonomic 

group. BDE 209 was detected in the larva and pupa of the Limnephilidae and Polycentropodidae 

families of caddisflies but not in the Ceratopogonidae and Chironomidae families of midge flies. 

Caddisfly pupa of the Polycentropodidae family had the highest overall level of BDE 209 

(4.93 ng/g dwt). BDE 47 was the most consistently-detected congener in larva and pupa of 

midges and caddisflies, with the exception of midge larva of the Chironomidae family, where it 

was not detected.

 Sellström et al. (2005) reported on the bioaccumulation of PBDEs from contaminated 

sewage sludge-amended soils. Earthworms were collected from five sites in Sweden. Three 

sites were government-operated experimental agricultural research workstations where the land 

application of sewage sludge was carefully controlled. The other two sites were private farms. 

One farm accepted the land application of sewage sludge; the other did not. Sewage sludge had 

been applied to the land at these sites as soil amendment from 1978−1997. The sludge applied to 

private farmland-received sewage sludge from a local STP. The STP also accepted and treated 

wastewater from a local textile industry that used PBDEs in their processes. Triplicate soil 

samples were taken from each site. Earthworms were collected from the same areas as soil

samples. Soil and earthworm samples were sent to a laboratory for chemical analysis using gas 

chromatography coupled with mass spectrometry run in the chemical ionization mode. In soil 

samples, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, and BDE 154 were analyzed, 

and when reported, the total PBDE includes these congeners. In earthworm samples, BDE 35, 

BDE 47, BDE 49, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, 

BDE 196, BDE 197, BDE 198, BDE 203, BDE 206, BDE 207, and BDE 209 were analyzed, and 

these comprised total PBDEs when results were reported. The researchers examined the 

relationship between the application rates of sewage sludge and the resulting levels of PBDEs in 

amended soils. In general, they found that the higher the amounts of PBDE-contaminated sludge 

applied to land, the more the BDE concentrations increase in soils, including BDE 196, 

BDE 197, BDE 206, BDE 207, and BDE 209 (Sellström et al., 2005). BDE 209 concentrations 

in soils from the research workstations and a private farm ranged from 0.015 to 0.75 ng/g dwt. 

Earthworms ingest soils as they move through soil. When in contact with the 
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PBDE-contaminated soils, the earthworms accumulated all the BDE congeners analyzed. Total 

PBDEs in earthworms ranged in concentration of all sludge-amended sites from 33 to 

38,000 ng/g lwt. The highest whole worm concentration of total PBDEs was from the farm 

where sewage sludge contaminated with PBDEs from a textile industry was used as soil 

amendment. BDE 209 concentrations in earthworms ranged in mean concentration of from not 

detected to 5,200 ng/g lwt. Other congeners showed the following ranges in mean 

concentrations, in ng/g lwt: BDE 47: 1.4−10,000; BDE 85: not detected (ND): 290; BDE 99: 

14−13,000; BDE 153: 1.1−1,200; BDE 154: ND: 870; BDE 183: ND: 0.71; BDE 206: ND: 830; 

and BDE 207: ND: 150. 

3.8. ENVIRONMENTAL TIME TRENDS

Sediment cores, archived vegetation, and biological tissue samples have been used to infer

time trends of the levels of PBDEs in the environment. The dated sediment cores and archived 

vegetation provide a clear record of the varying PBDE concentration over the decades. Whale 

blubber, bird eggs, and bird fat have been studied at different time intervals and can be used to 

evaluate the time trends of PBDE levels in biological tissues. The following is a limited review 

of these studies; further detail can be found in the studies themselves. These studies all show a 

similar trend: that PBDEs were absent in the environment until their introduction as flame 

retardant products in the 1970s. However, several of the sediment core studies showed 

concentrations in cores dated prior to the 1970s. Their presence at these layers was likely due to 

sediment bioturbation, imprecise dating, or a similar explanation, since PBDEs were not in the 

environment at these early dates. After the 1970s, the presence of PBDEs in the environment 

increased throughout the remainder of the 20th century into the 21st century. 

3.8.1. Time Trends from Sediment Core Studies

 Sediment core studies have been used to study 20th century temporal trends in 

environmental levels of such contaminants as PCBs, dioxins, and PBDEs. The approach is to 

take several cores usually within a lake (which is quiescent unlike moving water bodies such as 

rivers), section the cores by taking slices, and date the slices using radiotracers. Once a section is 

dated, then a measurement of the contaminant in that segment provides an indication of relative 

environmental levels during that time frame. Further, knowing the deposition rate of sediments 
3-64

in that particular water body, the total loading into that water body can be estimated. This 

approach is useful for persistent contaminants that predominantly partition to benthic sediments, 

such as PBDEs. This section provides an overview of sediment core studies undertaken for 

PBDEs around the world and what has been learned from them. 

Li et al. (2006) reported on the chronology of deposition flux of PBDEs into the 

sediments of all five Great Lakes (Lake Michigan, Lake Huron, Lake Superior, Lake Ontario, 

and Lake Erie) and three inland seepage lakes. Their work was also chronicled in three literature 

articles (Song et al., 2005a, b; 2004). Twenty-two sediment cores were collected in 2001 and 

2002 and horizontally sectioned into a total of 247 samples. Analytical results are separately 

reported as the sum of nine PBDE (Σ9BDEs) congeners (BDE 28, BDE 47, BDE 66, BDE 85, 

BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183) and BDE 209. The range of mean 

surface sediment concentrations in all the Great Lakes was 1.4−5.6 ng/g dwt and 

10.5−226.6 ng/g dwt for the Σ9BDEs and BDE 209, respectively. BDE 209 dominated the total 

PBDE concentrations in all the lake sediments. Figure 3-1 graphically displays the temporal 

trends of deposition fluxes of total PBDE and PCBs to the sediments in each of the Great Lakes 

as determined by Li et al. (2006). In general, PBDE levels began to rise between 1920 and 1950. 

This is not a reflection of when use of PBDEs began, because they were manufactured in the 

1970s. Rather, it is likely the result of sediment bioturbation once deposition occurred, or cross 

contamination during the coring and slicing process. There is a striking increase in the 

deposition of total PBDE to the lake sediments from 1970s to 2002. The temporal trends in the 

dated sediment cores from this study suggest that the increase in PBDE input to the Great Lakes 

tend to be first order. From these data, Li et al. (2006) calculated concentration doubling times 

(t2) ranging from 9 to 43 years for Σ9BDEs and from 7 to >70 years for BDE 209. There is no 

evidence of any recent decline in PBDE loadings to the Great Lake sediments. Li et al. (2006) 

also estimated the annual PBDE loading rates to the surface sediments of the Great Lakes by 

multiplying the mean surface flux times the total surface area of the lakes. According to their 

calculations, the Great Lakes received approximately 0.17 tons of Σ9BDEs and 4.4 tons of 

BDE 209 in 2002, primarily from atmospheric deposition. Total deposition of BDE 209 to the 

Great Lakes was estimated to be over 25 times the deposition of the sum of BDE 28, BDE 47, 

BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183 combined. 
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 Qiu et al. (2007) evaluated the distribution of flame retardants Dechlorane Plus (DP), 

1,2-bis-(2,4,6-tribromophenoxy)ethane (TBE), and PBDEs in dated sediment layers in a 

sediment core study of Lake Ontario. They measured 20 BDE congeners, including BDE 209. 

Figure 3-1. Time trends of deposition flux of PBDEs (blue solid diamonds) and 

PCBs (red open diamonds) to the sediments in each of the Great Lakes. (Left x-axis 

shows PBDE levels, and right x-axis shows PCB levels).

Source: Li et al. (2006).

Notes: SU12 = 12 samples from Lake Superior; LM18 = 18 samples from Lake Michigan; 

HU12 = 12 samples from Lake Huron; ON 48 = 48 samples from Lake Ontario. The x-axis shows the year 

dated by Pb-210 in the sediment core. 

In the sediment surface corresponding to 2004, Qiu et al. (2007) detected Σ3-7PBDE congeners 

(i.e., tri thru hepta BDE congeners BDE 28, BDE 47, BDE 49, BDE 99, BDE 100, BDE 116, 

BDE 153, BDE 154, BDE 181, and BDE 183) at a concentration of 2.8 ng/g dwt. The 

concentration of BDE 209 was approximately 15 ng/g dwt in the surficial layer, which was 

approximately five times higher than the sum of the other BDE congeners. The authors 

postulated that the sedimentary record of PBDEs in Lake Ontario reflected atmospheric 

deposition as the primary transport mechanism to the lake. BDE 209 was initially detected in Deposition flux = ng/cm2 per year
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Lake Ontario sediments around 1980 and underwent a dramatic increase from 1990 to 2000. 

Σ3-7PBDE congeners were initially detected around 1955, gradually increased in concentration 

between 1960 and 1990, and sharply increased in concentration from 1990 through 2000. These 

early detections before 1960, as noted earlier, do not reflect use at that time, but rather perhaps 

sedimentation or cross contamination during coring and slicing. 

 In 2004, Zhu and Hites (2005) conducted core studies for 18 BDE congeners from two 

study sites: Lakes Michigan and Erie. Like the other Great Lakes studies, BDE 209 

predominated, making up 95−99% of the concentration. The concentration of BDE 209 is 

315 and 39 ng/g dwt in Lakes Michigan and Erie, respectively, while all other congeners 

measured 2.6 and 1.1 ng/g dwt in Lakes Michigan and Erie, respectively. This article plots the 

rise in concentration, showing how BDE 209 was about 10 ng/g dwt in 1960 in Lake Michigan 

and rose to above 300 ng/g dwt by the early 2000s. The sum of the others similarly started well 

below 1.0 ng/g in 1960 to end up at 2.6 by the 2000s. Similar trends were seen in Lake Erie, 

with 209 not showing up until 1980 (at a detection limit of 1 ng/g dwt) to rise to 38 ng/g dwt by 

the 2000s. 

Marvin et al. (2007) studied the temporal trends of PBDEs in archived freeze-dried 

samples of suspended sediment taken from the water column of the Niagara River feeding into 

Lake Ontario. The archive contained samples that had been collected from 1980 to 2002. The 

Niagara River flows into Lake Ontario, and there has been a history of heavy industry along its

river banks. A total of 16 BDE congeners were evaluated in the samples. PBDE concentrations 

in the suspended sediments significantly increased between the time frames 1980 and 1988. 

However, the most current samples indicate a decline in total PBDE concentration in suspended 

sediments since 2001. This is different than the sediment core studies above, which show 

increases―at least until the last dating time at 2004. This difference could be attributed to the 

differences in the type of samples taken (suspended particles versus dated sediment core layers) 

and the fact that BDE 209 was not specifically analyzed in the samples. Another factor that may 

contribute to this difference is that Marvin et al. (2007) indicated that the primary inputs to the 

Niagara River were industrial point sources in the watershed, whereas inputs of PBDEs to the 

Great Lake sediments were primarily from atmospheric deposition. 

 Chen et al. (2007b; with earlier discussions of the partial data set also in Mai et al., 2005) 

examined the temporal trends of three dated sediment cores obtained from the Pearl River delta 
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in China. A total of 17 congeners were measured in these cores. Generally, the total 

concentrations increased from 1975 to the late 1980s and early 1990s. BDE 209, in particular,

displayed an exponential increase in concentration in the sediments from the time frame 1990 to 

1995 and the year 2005. Chen et al. (2007b) surmised that the sediments reflect the dominant 

use of the pentaBDE formulation prior to 1990, and a subsequent dominance of the use of 

decaBDE from 1990 onward. Deposition flux of BDE congeners, not including BDE 209, 

ranged from 5.8−106.2 ng/cm2, with an average of 56.0 ng/cm2, while the deposition flux of 

BDE 209 was higher, ranging from 172.8 to 563 ng/cm2. Chen et al. (2007b) estimated the 

deposition flux of total congeners, not including BDE 209, to the Peal River delta in 2005, to be 

2.1 and 29.7 metric tons (MT), respectively. 

 Minh et al. (2007) analyzed the occurrence and distribution of 11 BDE congeners, 

including BDE 209, in three dated sediment cores of Tokyo Bay. The investigators reported 

results in terms of the sum of BDE 3 through BDE 207 (ΣPBDEs) and BDE 209, separately. 

There were differences in the chronology of mass concentration of PBDEs from one sediment 

core to another. In addition, there was a clear concentration gradient of ΣPBDEs and BDE 209 

in the surface sediments, with the sediments at the mouth of Tokyo Bay having the highest 

concentrations. The authors indicated that the apparent concentration gradient clearly 

demonstrated that populated areas such as the cities of Tokyo and Yokohama are major emission 

sources of PBDEs to the Bay. Core dating data suggested that the ΣPBDEs concentrations 

consistently increased beginning in 1945 and reached a maximum in 1988, with concentrations at 

about 3 ng/g dwt or less at that maximum. There appeared to a slight decrease from 1988 to the 

surface sediments representative of the year 2000. BDE 209 appeared in the cores in the 1960s, 

reaching maximums between 20 and 80 ng/g dwt in 2000, the last year of dating. The authors 

explained that the apparent decease in the ΣPBDEs concentration in the mid 1990s may reflect 

Japan’s phase out and reduction in the general use of penta- and octaBDE commercial 

formulations in the early 1990s, and similarly, that the continual rise of BDE 209 reflects 

ongoing and increasing use of decaBDE.

Stern et al. (2005) reported the results of a sediment core study of a remote northern 

Arctic lake on Devon Island in Canada. A number of persistent organic pollutants, including 

PBDEs, were investigated in order to observe the temporal trends of contaminant loadings. 

Four sediment cores were collected in 1999 and archived in a freezer. Core #2 was analyzed for 
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the concentration of PBDEs. Maximum deposition of total PBDEs occurred in the most recent 

surface sediments with a depositional flux estimated at 28.5 ng/m2-year. BDE 47 was the most 

abundant congener, followed by BDE 99 and BDE 100. These three congeners represented 

about 80% of total PBDEs determined in the sediment core slices. The authors speculated that 

this congener pattern may be indicative of anaerobic microbial decomposition of more highly 

brominated PBDEs in the sediments, although there was no direct proof that this had occurred 

(Stern et al., 2005). The authors concluded that the contamination of the remote Arctic lake with 

PBDEs was the result of the long-range transport and deposition into the lake, because there 

were no local sources of these contaminants. 

Evenset et al. (2007) undertook a sediment core study to examine the historical 

deposition of PBDEs and other persistent organic pollutants into the sediments of Lake 

Ellasjoen, a remote lake on an island in the central Barents Sea of the Norwegian Arctic. 

Four replicate sediment cores were collected in April 2001 from a depth of 34 m and measured 

for 10 BDE congeners (not including BDE 209). PBDEs could only be detected in the upper 

4 cm of the sediment core, which corresponds to a time range encompassing the 1940s through 

2001. Of the 10 BDEs, only BDE 28, BDE 47, BDE 99, BDE 100, and BDE 153 were detected. 

They were first detected in core segments corresponding to 1953, at levels of 0.1 ng/g dwt and 

less, and they continuously rise in concentration to 1994, with the maximum individual congener 

concentration of 0.45 ng/g dwt. This dominant congener was BDE 47, followed by BDE 99, 

BDE 28, and BDE 100. BDE 153 was only found in the surface sediments to the lake. The 

primary route of entry of BDE congeners was assumed to be from atmospheric deposition, thus 

indicating the long-range transport of PBDEs (Evenset et al., 2007).

 Zegers et al. (2003) summarized various studies on the chronology of PBDE 

concentrations in dated sediment cores collected in Europe. The sediment core samples were 

obtained from the Oslofjord River in Norway; from the marine sediments of the Wadden Sea off 

the coast of The Netherlands; from Lake Woserin in the state of Mecklenburg-Vorpommern, 

Germany; and from the Kimmeridge clay formation in the United Kingdom, a marine formation 

from the Jurassic period. Pb-210 and Cs-137 (one or both) isotopes were used to mark distinct 

years in the age of the sediment layers. The findings of all of these studies are similar to all of 

the previously described core studies: Zegers at al. (2003) showed rising concentrations from 

about the 1970s to the present, with dominance of BDE 209, particularly in the later years. From
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the Drammenfjord River in Norway, BDE 47 initially appeared in the sediments in 1975 at 

1.2 ng/g dwt and steadily increased to a level of 5.8 ng/g dwt in 1999. Similarly, BDE 209 also 

first appeared in the sediments in 1975, but it increased at a sharper rate than BDE 47. The 

surface sediment concentration of BDE 209 is 80-fold higher than in 1975, rising from 1.3 to 

105 ng/g dwt. The concentration pattern from high to low suggests that BDE 209 > BDE 99 

> BDE 47 > BDE 100. In Lake Woserin in Germany, the oldest sediment layer was dated to 

1628, and the most current was 1997. BDE congeners initially appeared in 1973, continued to 

gradually increase, and reached a peak in 1994, with slight decreases seen for all congeners from 

1994 to 1997. BDE 99 was the most abundant congener in the sediments and increased threefold 

in concentration from 1973 to 1994 to a high of 11.3 ng/g dwt. Unlike other study results, 

BDE 209 did not dominate the profile. Rather, it had concentrations similar to BDE 99 and 

BDE 47, reaching a high of 10.7 ng/g dwt in 1994. Clay layers from the Kimmeridge Clay 

Formation (Blackstone-Band) in southern England, dated from the Jurassic period 100,000 to 

150,000 year ago, showed no detectable PBDE congeners. BDE congeners initially appeared in 

cores dated 1965 from the western Wadden Sea off the coast of the Netherlands, although the 

dating resolution of decades was not sufficient to determine if PBDEs were present from 1945 to 

1965. All detectable BDE congeners increased in concentration from their initial appearance 

until 1989. From 1989 to 1995, all congeners slightly decreased in sediment concentration. 

BDE 209 was the most abundant congener in the sediments from 1978 through 1995, reaching a 

high concentration of 380 ng/g dwt. In 1995, BDE 47 and BDE 99 reached high 

concentrations―about 20 ng/g dwt. Concentrations of 12 other congeners were mostly 

nondetects (at detection limits of about 1−3 ng/g dwt). 

 Many of the aforementioned sediment core studies suggest the presence of PBDEs in lake 

sediments prior to 1976 when PBDEs were intentionally manufactured (IPCS, 1994). This 

suggests that the core slices may not be precisely dated (relative dates rather than actual dates), 

or that mixing (bioturbation) from more recent sediment layers to lower layers may have 

occurred. However, when considered in their entirety, the sediment core studies do indicate the 

anthropogenic contribution of PBDE emissions to the environment. 
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3.8.2. Time Trends from Aquatic Wildlife Samples

The body burden of PBDEs and other persistent organic pollutants in whales, seals, and 

different fish species has been measured over discreet time intervals. This provides a basis for 

assessing the changes in relative body burdens in wildlife with the passage of time. Like 

sediment core studies, these studies show a rise of BDE congeners through the 1990s into the 

2000s. However, unlike sediment core studies, these show a predominance of BDE 47 and a 

virtual absence of BDE 209. 

Lebeuf et al. (2004) studied the levels and temporal trends of PBDEs in the blubber of 

54 stranded adult beluga whales from the St. Lawrence Estuary in Quebec, Canada. The 

beaching of whales on the shores of the St. Lawrence Estuary had occurred in the time frame

1988 through 1999, and samples were collected during these years and frozen for future analysis. 

The total PBDE concentration (including BDE 28, BDE 47, BDE 71, BDE 77, BDE 99, 

BDE 100, BDE 153, BDE 154, BDE 155, and BDE 183) ranged from about 20 to almost 

1,000 ng/g wwt in samples from 54 whales. The data suggest an exponential increase of PBDE 

congeners from 1988 to 1999. The authors estimate the doubling times of BDE congeners to 

range between 2 and 9 years for all congeners (Lebeuf et al., 2004).

Kajiwara et al. (2004) used archived fur seal adipose tissue samples to investigate the 

relative time-trends of PBDEs in coastal waters of Japan. Ten fat samples had been collected 

over a period between 1972 and 1998 and stored in the Environmental Specimen Bank for 

Global Monitoring at Ehime University. Kajiwara et al. (2004) analyzed tissue samples for the 

presence of PCBs, DDT, and BDEs (BDE 3, BDE 15, BDE 28, BDE 47, BDE 99, BDE 153, 

BDE 154, BDE 183, and BDE 209). The sum of the PBDE congeners ranged in concentration 

from a low of 0.33 ng/g lwt in the year 1972 up to 100 ng/g lwt in the year 1994. BDE 47 was 

the most abundant congener of the total PBDEs in all samples analyzed. No BDE 209 was 

detected in the fur seal fat samples despite the fact that the decaBDE formulation constituted 

67% and 100% of total commercial PBDE usage in Japan in 1985 and 2000, respectively.

Batterman et al. (2007) reported on the time-trends of PBDEs in archived frozen tissues 

samples from rainbow smelt, walleye, and lake trout obtained from the Great Lakes during the 

time frame from 1979−2005. All fish species had been collected from sites in Lakes Erie, 

Huron, Michigan, Ontario, and Superior every other year as part of a monitoring program 

conducted by EPA. Only BDE congeners BDE 47, BDE 99, BDE 100, and BDE 153 were 
3-71

investigated in this study due to their usual pattern of dominance in fish tissue. The total PBDE 

concentrations in trout at each of the lakes increased exponentially and rapidly over the period 

from 1979−1980 to the mid-1990s. Concentrations were always highest for BDE 47, rising to 

greater than 100 ng/g wwt in the latest samples in all lakes, while concentrations of the other 

three congeners were between 5 and 10 times lower than BDE 47, rising to no more than 

15 ng/g wwt. Doubling times were calculated for all species and lakes. They ranged from about 

2 to over 20 years. The relatively stable congener pattern in all lakes across all fish species (i.e., 

BDE 47 > PBE 100 > PBE 99 > PBE 153) suggests that atmospheric deposition was the primary 

reason for the presence of PBDEs in the Great Lakes. The authors claim that there is no 

evidence to suggest any declines in PBDE concentrations in the Great Lakes region.

 Rayne et al. (2003b) reported on the time trends of PBDEs in tissues of 41 mountain 

whitefish and 6 sucker fish in the Columbia River system in southeastern British Columbia, 

Canada. Eleven congeners were measured but not BDE 183 and BDE 209. From 1992 to 2000, 

the ΣPBDE congeners in whitefish increased 11.8-fold and 6.5-fold in fish caught near the towns 

of Genelle and Beaver Creek along the Columbia River, respectively. The authors calculated a 

short doubling time of total PBDEs of 1.6 years between the years 1995 and 2000 in whitefish 

caught near Genelle. At the confluence of Beaver Creek and the Columbia River (25 miles 

further downstream from Genelle and 9 miles downstream of a secondary metal smelting 

operation), the PBDE concentrations increased over six fold from 1992 to 2000―from 4.5 ng/g 

wwt in 1992 to 29.2 ng/g wwt in 2000. In whitefish caught on the Slocan River in an 

unpopulated pristine area that was not directly impacted by urban or industrial activities, PBDE 

concentrations were 0.9 ng/g wwt in 1996, which is about 20−50 times lower than those at other 

testing locations near towns and possible sources of PBDEs. Based on concentration differences 

in species, the authors speculated that PBDEs bioaccumulated less in sucker fish than whitefish. 

3.9. CONCLUSIONS

 This chapter focused on the environmental fate of PBDEs. Sections have discussed fate 

properties, movement, and transformations in the environment, temporal trends in the 

environment, and also evidence for metabolic transformations, specifically debromination, in 

animals and humans. The following conclusions are made: 
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1. In general, PBDE congeners are lipophilic and persistent organic compounds having a 

propensity for bioaccumulation and biomagnification in the aquatic and terrestrial food 

webs. BDE 209 is a large molecule having reduced tendency for bioaccumulation. 

However, BDE 209 has been measured in tissues of terrestrial birds and mammals, and 

aquatic fish and mammals. Therefore, BDE 209 is contaminating the aquatic and 

terrestrial food chains. 

2. PBDE congeners are ubiquitous environmental contaminants and are detected globally in 

air, soils, sediments, oceans, and wildlife. 

3. The atmospheric transport and surface deposition of PBDEs is the primary means of 

distributing PBDEs over long geographical distances. 

4. The detection of BDE congeners in arctic air suggests long-range atmospheric transport 

of these contaminants from industrialized countries. 

5. Once released into the air, PBDEs partition between the vapor and particle phases in the 

atmosphere in accordance with their respective vapor pressures. The BDE congeners 

with 1−4 bromine atoms primarily exist in the vapor phase; BDE congeners with 

5−6 bromine atoms more generally partition between the vapor phase and the particle

phase, and BDE congeners having >6 bromine atoms are primarily adsorbed to 

atmospheric particles.

6. Photochemical reaction with the hydroxyl radical appears to be an insignificant 

atmospheric degradation pathway for vapor phase BDE congeners. 

7. High molecular weight BDE congeners exhibit a propensity for the breakdown and 

decomposition in soils by UV light (i.e., photolysis). Photo-degradation in soils can 

debrominate the higher BDE congeners to form lower-brominated BDE congeners and 

may be a significant degradation pathway in the environment. Photolysis of BDE 47 and 

BDE 99 in air is a major atmospheric removal process for these congeners. In contrast, 

the photolysis of BDE 209 in air is a minor atmospheric removal process. Atmospheric 

wet and dry deposition of BDE 209 is the most significant atmospheric removal pathway. 

This is likely due to the fact that BDE 209 predominantly exists bound to particles in air. 

8. Higher-brominated PBDE congeners can undergo metabolic debromination in fish, 

mammals, and birds to form lower-brominated congeners. 

9. There exists good evidence that microbial anaerobic degradation naturally occurs in 

sediments and soils. However, the degradation rate has yet to be determined. 

10. Soils and sediments are environmental sinks for PBDEs. 

11. Since the 1970s, sediment core samples show predominance and a starker rise of 

BDE 209, while the animal tissue samples show predominance and rise in BDE 47. 

12. The environmental fate of BDE congeners is dependent on their respective chemical and 

physical properties. 
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13. A principal source of PBDE in indoor microenvironments is household and commercial 

dust. PBDEs volatilize from PBDE-treated products to indoor dust, and or transfer to 

dust from the abrasion of products. 

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4. ENVIRONMENTAL AND EXPOSURE MEDIA CONCENTRATIONS

4.1. INTRODUCTION

This chapter summarizes the concentrations of individual congeners and total 

polybrominated diphenyl ethers (PBDEs) in environmental and exposure media. The emphasis 

is on data from the United States, although key data sets from other countries will be described 

as a contrast and a supplement. In addition, emphasis will be placed on the measurements for 

brominated diphenyl ether (BDE) 209, as this deca congener is essentially the sole component of 

the deca formulation that is still currently being produced and marketed in the United States. 

Much of the literature has focused on the profile of congeners (“profile” is the term used in this 

chapter to describe the suite of congeners) associated with the penta formulation (BDE 47, 

BDE 99, BDE 100, BDE 153, and BDE 154), and, subsequently, there is a paucity of data for the 

primary markers of the octa formulation, BDE 183 (about 40% of the octa formulation), and the 

deca formulation, BDE 209 (about 98%). BDE 183 is often considered a marker for the octa 

formulation because it is a primary congener found in this formulation, while it is not found in 

the penta or deca formulations. Also, it is persistent in the environment and often found at low 

levels when measured. There is suggestive evidence that photolytic debromination of BDE 209 

results in the formation of BDE 183 (Stapleton and Dodder, 2008), so its presence could also 

reflect past use of the deca formulation. As the deca formulation is the primary formulation of 

PBDEs marketed and used worldwide, BDE 209 is now more regularly measured, but that was 

not always the case. Surveys in the 1990s into the 2000s mostly did not include BDE 209, and 

even current surveys often do not include this critical congener. Analytical difficulties, evidence 

that it is rapidly eliminated from biota, and citation of previous literature focusing on the 

prevalence of the penta formulation congeners, are the primary reasons given for not measuring 

or reporting BDE 209. While nearly all studies include at least the five key congeners of the 

penta formulation, some studies report on the concentrations of over 20 congeners. Unlike 

dioxin-like compounds where 7 dioxins and 10 furans have been identified as toxic congeners of 

concern, there is no list of key toxic BDE congeners. Subsequently, there is not uniformity on 

which congeners should be measured, and when citing a “total” concentration (sum of all 

congeners) measured in a study, one must be careful to identify which congeners constitute that 

total. Efforts were made in the discussions below to indicate which congeners were measured in 
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a study, and when referencing “total” for each study, it is only the total of the congeners 

measured. The original references should be obtained if the discussions below are not clear. 

The sections on air, house dust, fish, and food include a comprehensive table of reported 

measurements of individual congeners in the United States; sections on water, sediment, and 

surface soil do not include such tables because there are little or no U.S. data on these media in 

the literature. The chapter concludes with development of a tabular assignment of BDE 

congener concentrations for each exposure media. These assignments will be used in the next 

chapter, which determines exposure dose of BDEs based on exposure media concentrations and 

contact rates. 

4.2. WATER AND SEDIMENT CONCENTRATIONS

Water has been rarely sampled for PBDEs in America, and it is questionable whether 

available monitoring can be considered representative of drinking water. The San Francisco (SF)

Estuary Regional Monitoring Program for Trace Substances sampled water, surface sediments, 

and bivalves (oysters, mussels, and clams) in the SF Estuary for 22 BDE congeners, including 

BDE 209 (Oros et al., 2005). Thirty three water samples were taken, with total PBDEs ranging 

from 3−513 pg/L, and a mean concentration of 146.2 pg/L. The region of the bay that had the 

highest concentrations, the Lower South Bay, receives 26% of the Estuary’s total publicly owned 

treatment works wastewater effluents and only 10% of the Estuary’s freshwater inflow. The 

most abundant congeners were BDE 47, BDE 99, and BDE 209. When BDE 209 was not 

reported as “Q” (outside QA limits, detected but not reported), it was present and quantified, and 

of a comparable magnitude as BDE 47, though mostly at a slightly higher concentration. 

BDE 47 concentrations ranged from about 17 to greater than 60 pg/L, with one high sample at 

123 pg/L; BDE 99 was next highest with quantified concentrations (about 1/3 of the samples 

were outside quality limits) ranging from about 11 to 35 pg/L, with one high level at 90 pg/L. 

BDE 17, BDE 100, and BDE 28/33 (congeners notated as “28/33” translate to congeners that 

have coeluted and are analyzed as though they were one congener) were quantified at levels less 

than 5 pg/L, and others not below detection or very infrequently detected at levels less than 

5 pg/L. It was found that the PBDEs were predominantly associated with the sediment fraction

of the water column. 
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 In a second study, Johnson et al. (2006), sampled for PBDEs in water and fish of 

Washington state rivers and lakes. Results from 15 samples taken in seven rivers and three lakes 

in 2005 and 2006 were available. Congeners measured included BDE 47, BDE 49, BDE 66, 

BDE 71, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, BDE 184, BDE 190, and 

BDE 209, although BDE 209 was never detected. Total concentrations ranged from 1 to 

926 pg/L, although only two samples were above 100 pg/L, with one of them at 146 pg/L. The 

average was 91 pg/L, but there was a much lower median of 16 pg/L. The most abundant 

congeners found, and at the highest concentrations, were BDE 47, BDE 99, and BDE 100. 

 Finally, Streets et al. (2006) measured PBDEs in the dissolved and particle phases in 

Lake Michigan and found BDE 47, BDE 99, BDE 100, and BDE 66 in the dissolved phase in 

concentrations ranging from 0.13 to 10 pg/L (for individual congeners). The quantified particle 

phase concentrations for BDE 47, BDE 99, and BDE 100 ranged from 0.18 to 1.4 pg/L. 

BDE 209 was not evaluated in this study. 

 In summary, only three studies measuring surface water concentrations in the United 

States could be found. In one study, total concentrations in the SF estuary ranged from 3 to 

513 pg/L, dominated by BDE 47, BDE 99, and BDE 209. In the second study, BDE 209 could 

not be quantified, and total concentrations of the lower-brominated congeners ranged widely 

from 1 to 916 pg/L. Although only two samples were above 100 pg/L, and one of them was at 

146 pg/L. The mean and median in that study were 91 and 16 pg/L, respectively. The third 

study did not measure BDE 209, and, in that study, the authors separated out the dissolved and 

particle phase concentrations. In total, only the sum of BDE 47, BDE 66, BDE 99, and BDE 100 

ranged up to10 pg/L. 

Much of the data on PBDEs in sediments of water bodies were taken in the context of 

sediment core studies, whose purpose is to elaborate on temporal trends of PBDEs in the 

environment (see Section 3.8.1). This section provides an overview of studies just focusing on 

surface sediments and on the surficial sediment concentrations in the sediment core studies. 

Briefly, the sediment core studies showed BDE 209 to dominate surficial sediment 

profiles. The average concentrations of BDE congeners, not including BDE 209, in these studies 

were generally less than 10 ng/g, while BDE 209 concentrations ranged anywhere from 10 ng/g 

to 300+ ng/g. The maximum total BDE concentrations found exceeded 1,000 ng/g in some 
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cases, and they were dominated by BDE 209, which comprised over 90% of the total mass 

measured. 

The study of the SF Estuary included both water and sediment sampling (Oros et al.,

2005). A total of 48 sediment samples were taken, and positive detections were only noted for 

BDE 47, BDE 99, BDE 183, BDE 204, and BDE 205. It is noteworthy that BDE 209 was not 

found with a detection limit (DL) of 1.5 ng/g. This contrasts with all other sediment studies 

where BDE 209 dominated the profile. It is also the only study found that provided 

concentrations of BDE 204 and BDE 205. Total concentrations ranged from ND (not detected) 

to 212 ng/g dry weight (dwt) basis. BDE 47 was detected 42% of the time, with a range of 

detections of 1.1 to 100 ng/g dwt and a mean of 12 ng/g dwt. BDE 99 was detected 77% of the 

time with a range and mean of 0.3 to 71 and 5 ng/g dwt, respectively. BDE 183, BDE 204, and 

BDE 205 were detected once, twice, and once, respectively. 

 Hun Yun et al. (2008) sampled PBDEs and polychlorinated biphenyls (PCBs) in 

sediment and flood plain soils of the Saginaw River Watershed in Michigan. They took 

53 samples representing 0−12 cm in October−November 2004. They sampled for 10 congeners: 

BDE 28, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, and 

BDE 209. The total concentration (sum of these congeners) ranged from 0.04−49.9 ng/g dwt, 

with an average of 3.75 ng/g dwt. BDE 209 dominated the concentration, explaining 3.19 ng/g 

dwt of this average. 

 Song et al. (2005a, b; 2004) collected 22 sediment cores in 2001 and 2002 and 

horizontally sectioned them into a total of 247 samples. Analytical results were separately 

reported as the sum of nine PBDE (Σ9BDEs) congeners (BDE 28, BDE 47, BDE 66, BDE 85, 

BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183) and BDE 209. The range of mean 

surface sediment concentrations in all the Great Lakes was 1.4−5.6 ng/g dwt and 

10.5−226.6 ng/g dwt for the Σ9BDEs and BDE 209, respectively. BDE 209 dominated the total 

PBDE concentrations in all the lake sediments. 

Raff and Hites (2004) collected suspended sediment samples from 26 sites along the 

Mississippi River and five of its major tributaries during July−August 2002 and March 2003. A 

total of 15 congeners were measured. Individual congeners were not identified nor were 

individual congener concentrations provided. Total concentrations ranged from 

31−1,548 ng/g dwt, with an average of 327 ng/g dwt. Consistent with the Great Lakes sediment 
4-5

cores described in Chapter 3, BDE 209 was the overwhelmingly dominant congener. On 

average, it comprised 96.8% of the total concentration. Congeners 47 and 99 were the only other 

congeners accounting for concentrations in excess of 1% of the total PBDE 

concentrations―1.16% and 1.26%, respectively. Based on concentrations near the mouth of the 

river, combined with records on suspended sediment concentrations and outflows, the authors 

estimated that 8 tons/year of PBDEs are discharged into the Gulf of Mexico. Other interesting 

trends include higher concentrations in the spring, attributed to high runoff, and evidence of 

debromination, as two sites contained higher concentrations of the nona congener BDE 206, as 

compared to BDE 209. 

Ashley et al. (2006) analyzed four sediment samples, along with samples from eels taken 

from the Delaware River (eel data summarized in Section 4.6.2). The sediment samples were 

collected in 2002 as part of an earlier PCB study, and they were reanalyzed in 2006 for PBDEs. 

While the total concentrations in these samples were lower than some of the studies 

above―between 0.7 and 21.7 ng/g dwt―the proportions were similar. BDE 209 explained 

about 50% of the concentration, while BDE 99 and BDE 47 were next, accounting for 15% and 

14%, respectively. 

Four surficial sediment samples were taken in Lake Hadley in Indiana (Dodder et al., 

2002). Total BDE concentrations, including BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, 

and BDE 209, ranged from 24 to 71 ng/g dwt, dominated by BDE 209 at 19 to 33 ng/g dwt. 

Other than a measurement of 22 ng/g dwt for BDE 99, all other measurements of the other 

congeners were near or less than 5 ng/g dwt. 

 Hale et al. (2002) studied the environmental impacts of a polyurethane foam (PUF) 

manufacturing facility in the U.S. mid-Atlantic Region, which had ceased production in 1997; 

sampling occurred in 2001. An interior duct of the facility was tested, as was soil adjacent to the 

facility, sediment in a stream leaving the facility, and sediment and bluegill in a pond about 

250 m from the facility. Total concentrations of PBDEs (including BDE 47, BDE 99, and 

BDE 100; BDE 153, and BDE 154 were measured but not detected) in surface sediments in the 

stream leaving the facility were 17.2 ng/g dwt and 132 ng/g dwt, with one sample having all 

nondetects. Total concentrations in surface sediments were 0.5 and ND ng/g in two pond 

sediment samples. These concentrations are similar to measurements described at other settings, 
4-6

suggesting that the foam production facility did not result in increases to sediment concentrations 

compared to nearby aquatic settings. 

Hale et al. (2001) also measured PBDE (along with PCB) concentrations in fish and 

sediment samples from two large Virginia watersheds. The study sites included the largest 

bodies of freshwater in Virginia; they were not selected based on concerns from a particular 

source of contamination, although the general predominance of furniture manufacturing facilities 

was noted. Total PBDE concentrations, including BDE 47, BDE 99, BDE 100, BDE 153, and 

BDE 154 ranged from nondetect to 52.3 ng/g dwt in 17 samples. BDE 49 was also measured but 

not detected in sediment samples. The concentrations were dominated by BDE 47, explaining 

about 53% of total concentrations, with BDE 99 second, at about 35% of total, and the remaining 

concentration due to BDE 100 (7%), BDE 153 (<5%), and BDE 154 (<5%). 

 Much higher concentrations have been identified in sediments directly downstream of a 

known source or in heavily industrialized areas. La Guardia et al. (2007) reported on sediment 

concentrations downstream of a wastewater treatment plant of a plastics manufacturer in North 

Carolina. In eight sample points downstream of the outfall, four each in 2002 and 2005, they 

found total PBDEs (sum of 19 congeners, including BDE 47, BDE 99, BDE 100, BDE 153, 

BDE 183, BDE 203, BDE 206, BDE 207, BDE 209, and others) to range from about 300 to over 

3,000 parts per million (ppm; μg/g), with BDE 209 comprising over 95% of the total. This is 

over 4 orders of magnitude higher than reported above for background settings. They also took 

two samples upstream, in 2002 and 2005, and found much lower but still quite elevated 

concentrations of total PBDEs, at 38 and 97 ppm. 

Toms et al. (2006) undertook a comprehensive assessment of the fate and distribution of 

26 BDE congeners in the aquatic environment of Australia in 2003−2004. An aim of this study 

was to determine the background concentrations and congener compositions of BDEs in 

estuarine, freshwater, and marine sediments. Ninety sediment samples were analyzed from 

locations representing various land uses ranging from remote to industrial. A total of 25 BDE 

congeners were detected in samples from 35 of 46 sites (76%), and total PBDE concentrations 

ranged from nondetect to 60.9 ng/g dwt, with an overall mean (± standard deviation) and median 

of 4.7 ±12.6 and 0.3 ng/g dwt, respectively. BDE 209 dominated the congener distribution in 

86% of the sediment samples. 
4-7

Christensen and Platz (2001) sampled sediment from Danish marine coastal areas, 

freshwater lakes, and a river in 2000. The congeners measured included BDE 47, BDE 99, 

BDE 100, BDE 153, and BDE 209, and the total concentrations ranged from 0.06−24.7 and 

0.07−10.6 ng/g dwt in marine and freshwater sediment, respectively. BDE 209 dominated the 

congener profile in marine and freshwater sediments, with a median concentration of 

3.35 ng/g dwt and 2.05 ng/g dwt, respectively. The rank order of BDE congeners by 

concentration in sediment was BDE 209 > BDE 99 > BDE 47 > BDE 100 > BDE 153. The 

highest concentrations of BDEs were detected in sediments in harbors and lakes located in urban 

areas. Because PBDEs were never produced in Denmark, the authors postulated that PBDEs 

entered the Danish environment primarily by long-range transport. The authors suggested that 

additional local sources could be evaporation and leaching from PBDE-treated products. 

 In a study by Eljarrat et al. (2005), 13 marine sediment samples were collected from three 

coastal areas of Spain in 2002. The sediments were analyzed for 40 BDE congeners, but only 12 

were detected in the sediments: BDE 28, BDE 33, BDE 47, BDE 66, BDE 77, BDE 100, 

BDE 99, BDE 118, BDE 154, BDE 153, BDE 183, and BDE 209. Total PBDE levels ranged 

from 2.7 to 134 ng/g dwt in coastal marine sediments, with the highest levels of sediments off 

the coast of Barcelona. All the sediment samples were dominated by BDE 209, which 

constituted between 50 and 99% of the total PBDE contamination. The usual congener profile 

found in the coastal marine sediments was BDE 209 > BDE 47 > BDE 99 > BDE 100 

> BDE 153. 

Water body sediment studies in North America included one in the San Francisco 

Estuary, along the Mississippi and Delaware Rivers, one in a Virginia watershed, one in a 

Michigan watershed, one in a pond near a closed polyurethane foam manufacturing facility in 

North Carolina, and several surficial sediment and core studies describing trends in the Great 

Lakes (some of these studies are reviewed in Chapter 3). Overall, total concentrations were 

mostly less than 20 ng/g dwt but ranged as high as 300 ng/g dwt in Lakes Michigan and Erie. 

With one exception, the study in the SF Bay Estuary (Oros et al., 2005), over 90% of the total 

concentration was BDE 209 when it was measured. In the SF Bay, BDE 209 was not detected in 

sediments at all (detection limit sufficiently low at 1.5 ng/g dwt), and BDE 47 dominated the 

profile. In a few studies measuring BDE congeners in industrial settings and near outfalls of 

wastewater treatment plants of facilities producing PBDEs, substantially higher concentrations in 
4-8

the hundreds of μg/g (ppm) range are found, with over 90% of the total concentrations being 

BDE 209. 

4.3. SURFACE SOIL CONCENTRATIONS

 Only two systematic studies could be found that looked at PBDEs in surface soils in the 

United States (Offenberg et al., 2006; Hun Yun et al., 2008) in predominantly suburban and/or 

background settings. Offenberg et al. (2006) took a total of 33 surface soil samples in 15 states 

and measured for 30 BDE congeners. Concentrations of total BDEs averaged 103 ng/g and had 

a geometric mean concentration of 5.3 ng/g dwt and a range of 0.09 to 1200 ng/g dwt. BDE 47 

was detected in 31 of 33 samples, averaging 1.9 ng/g dwt over the entire data set (ND = 0). 

BDE 99 was observed in 30 samples and averaged 3.6 ng/g dwt. BDE 209 was found in 

24 samples and averaged 15.3 ng/g dwt. The highest concentrations were found for BDE 183, 

but it was only found in three samples at concentrations, ranging from 121 to 562 ng/g dwt, so 

that the survey-wide average was 37.4 ng/g dwt. Hun Yun et al. (2008) took 26 samples in 

floodplain soils of the Saginaw River Watershed in Michigan in October−November 2004. They 

measured for 10 congeners (BDE 28, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 138, 

BDE 153, BDE 154, and BDE 209) to a depth of 2.5 cm. Total concentrations (sum of 

10 congeners) ranged from 0.02−55.1 ng/g dwt, with an average of 5.6 ng/g dwt. BDE 209 

dominated the profile, explaining 5.4 ng/g dwt of this average. 

Hale et al. (2002) measured surface soil near a polyurethane foam production facility that 

had been closed a few years earlier than sampling. They quantified concentrations of BDE 47, 

BDE 99, and BDE 100 (BDE 153 and BDE 154 measured but not detected) in three samples 

with total concentrations of ND (no congeners detected), 13.6, and 76.0 ng/g dwt. In these three

samples, BDE 47 and BDE 99 dominated, explaining nearly half of the total concentrations each. 

Two European studies looked at background soils, and one study was found evaluating 

soil concentrations near an electronics recycling facility in China. Hassanin et al. (2004) 

reported on sampling of 66 surface samples (0−5 cm), and for 38 of these, a paired subsurface 

sample (5−10 cm) was taken from grassland and woodland areas in the United Kingdom (UK) 

and Norway. These samples were collected in 1998, and 20 BDEs were measured, although 

BDE 209 was not measured. The congeners most routinely detected (85−100% of the time) in

the samples were BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183. Major trends 
4-9

include the following: United Kingdom concentrations were higher than Norway concentrations, 

woodland concentrations were higher than grassland concentrations, and subsurface

concentrations were significantly lower than surface soil concentrations. The total concentration 

of all BDEs ranged from 0.065 to 12.0 ng/g dwt. The results for grassland/United Kingdom, 

woodland/United Kingdom, and woodland/Norway were presented in terms of medians, 

minimum/maximum and percent detected. The selected range of median concentrations for 

these three groupings of surface soils were (pg/g dwt, with percent detected in parenthesis): 

BDE 17 (52−75%)―27, 28, 35; BDE 28 (14−76%)―17, 21, 29; BDE 47 (100%)―61, 490, 250; 

BDE 99 (92−95%)―280, 900, 360; BDE 100 (90−95%)―36, 110, 58; BDE 153 (54−95%)―72, 

210, 51; BDE 154 (90−100%)―22, 100, 42; and BDE 183 (54−100%)―26, 70, 25. The authors 

state that the percentage found in soil mirrored the technical penta product, Bromkal 70-5DE, 

with BDE 47 averaging 21% of total PBDE, BDE 99 averaging 40%, BDE 100 averaging 6%, 

BDE 153 averaging 8.7%, and BDE 154 averaging 4.4%. A major shortcoming of this surface 

soil study, however, was that BDE 209 was not measured. Given high measurements of 

BDE 209 in sediments, dust, and other soil studies, its lack of measurement in this broad ranging 

surface soil sampling study is unfortunate. 

 A second study looked at soil samples in a rural-urban transect across the West Midlands 

of the United Kingdom, which includes the major urban center of Birmingham. The purpose of 

the study was to ascertain whether there was a relationship between concentrations and 

characterization of the location of sample (urban, suburban, or rural). Soil concentrations were 

found to be consistent with earlier sampling in the United Kingdom (Hassanin et al., 2004). 

Total soil concentrations (which included BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and 

BDE 154) taken at a 0 to 5-cm depth for 10 sites ranged from 0.07 to 3.89 ng/g dwt. The 

maximum concentration of 3.89 ng/g was found in the city of Birmingham, while other 

concentrations were 0.84 ng/g dwt and less. The ratio of the two primary congeners found, 

BDE 47 to BDE 99, ranged from 0.51 to 0.85. 

 Cai and Jiang (2005) characterized soil concentrations near an electronics recycling 

facility in China. For the six soil samples, the congeners measured, and their average 

concentrations were as follows: BDE 3 presented as ND (monoBDE); BDE 15―0.76 ng/g dwt

(DiBDE); BDE 28―5.01 ng/g dwt (triBDE); BDE 47―217 ng/g dwt; BDE 99―552 ng/g dwt; 

BDE 139―35 ng/g dwt; BDE 153―74 ng/g dwt; BDE 154―49 ng/g dwt; and 
4-10

BDE 183―12.3 ng/g dwt. The total concentration is 945 ng/g dwt. The authors state that the 

isomer pattern of BDE 47, BDE 99, BDE 139, BDE 153, and BDE 154 is consistent with the 

pentaBDE formulation. No measurements of BDE 209 were made in this study. 

In summary, the sparse literature suggests background soil concentrations of total BDEs 

in the United States might average above 100 ng/g dwt, although a central tendency estimate 

(geometric mean or median) might be lower, at near 5 ng/g dwt. Total concentrations in soils at 

industrial sites, which might suggest a more direct source, could range as high as 1,000 ng/g dwt. 

The congener pattern noted in the studies on background soils is similar to the pentaBDE 

formulation. However, BDE 209 mostly has not been sampled in soil studies; in the one study 

where it was measured, it was found in 24 of 33 samples, averaging 15.3 ng/g dwt. 

4.4. DUST CONCENTRATIONS

4.4.1. House Dust Concentrations

House dust was the focus of several U.S. studies because of the concern for indoor 

exposures in residences. Data were also available in a study on concentrations in a computer 

laboratory. As will be shown, house dust from both homes and places of work are dominated by 

BDE 47, BDE 99, and BDE 209. Table 4-1 provides congener-specific concentrations for 

U.S. house dust from the literature.

House dust and dryer lint samples were collected from 16 homes in the Washington, DC 

area and 1 from Charleston, SC. The samples were analyzed for 22 individual PBDE congeners 

(Stapleton et al., 2005). Total concentration in house dust ranged from 780 ng/g dwt to 

30,000 ng/g dwt, with a mean total of 5,900 ng/g dwt. The dominant congeners were the ones 

associated with the penta- and decaBDE commercial mixtures, BDE 47, BDE 99, and BDE 209. 

The mean concentrations of these three congeners were 1,220, 1,700, and 2,090 ng/g dwt, 

respectively. No correlations were found in the study with year of construction, type of flooring 

(hardwood vs. carpet), or the number of televisions and personal computers (PCs; and hours of 

computer use per week). However, an inverse relationship was found with the area of the home 

and the contribution of BDE 209 to the total PBDE concentration in dust (i.e., the larger the 

home, the smaller the BDE 209 concentration). Clothes dryer lint, examined in five of the 

homes, showed concentrations ranging from 480 to 3,080 ng/g dwt. The two house dust 
4-11

Table 4-1. Congener-specific concentrations of PBDEs in house dust in the 

United States (units in ng/g dwt) 

Congener Concentration

(ng/g dwt) Comment Reference

DiBDE

15 11 Mean 10 homes throughout U.S., but 9 ND, 1 at 

109

Sharp and 

Lunder (2004)

TriBDE

17 9 Mean 17 homes, Washington, DC Stapleton et al. 

(2005)

5.5 Carpet dust (n = 2) from computer labs in CA CARB (2005)

2.4, 6.4 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

1.4, 0.6, 0.4 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom., 

and vacuum

Allen et al. 

(2008)

25 2 Carpet dust (n = 2) from computer labs in CA CARB (2005)

28 21 Mean 17 homes, Washington, DC, also listed as 

BDE Congener 33

Stapleton et al. 

(2005)

ND (50) Mean 10 homes throughout U.S. Sharp and 

Lunder (2004)

14.3 Carpet dust (n = 2) from computer labs in CA CARB (2005)

3, 20.3 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

25, 14 Mean, geometric mean from homes in 

Amarillo/Austin, Texas from 2006 (n = 20)

Harrad et al. 

(2008b) 

0.25, <0.1 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

30 ND Carpet dust (n = 2) from computer labs in CA CARB (2005)

32 ND Carpet dust (n = 2) from computer labs in CA CARB (2005)

33 ND (50) Mean 10 homes throughout United States Sharp and 

Lunder (2004)

28/33 16.3, 10.5, 6.4 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

35 ND Carpet dust (n = 2) from computer labs in CA CARB (2005)

37 ND Carpet dust (n = 2) from computer labs in CA CARB (2005)
4-12

Table 4-1. Congener specific concentrations of PBDEs in house dust 

in the United States (units in ng/g dwt) (continued)

Congener Concentration

(ng/g dwt) Comment Reference

TetraBDE

47 1,220 Mean 17 homes, Washington, DC Stapleton et al. 

(2005)

1857 Mean 10 homes throughout U.S. Sharp and 

Lunder (2004)

ND 

(400)―9,860

Range of 89 homes sampled in Cape Cod, MA; 

45% detected

Rudel et al. 

(2003)

430 

(230−3,000)

Median, range from 10 homes in Atlanta Sjodin et al. 

(2008)

456 Carpet dust (n = 2) from computer labs in CA CARB (2005)

364, 1621 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

1865, 837, 338 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

669, 14613 Median, maximum from 11 homes in the Boston, 

MA area

Wu et al. (2007)

810, 470 Mean, geometric mean from homes in 

Amarillo/Austin, Texas from 2006 (n = 20)

Harrad et al. 

(2008b)

3750 Median, n = 49 homes in Northern CA Zota et al. (2008)

2920; 

655−8030

Median, range, n = 11 homes in Davis, CA Hwang et al. 

(2008)

93, 40 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

49 ND Carpet dust (n = 2) from computer labs in CA CARB (2005)

29.6, 23.6, 

12.4

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

66 28.5 Mean 17 homes, Washington, DC Stapleton et al. 

(2005)

21 Mean 10 homes throughout U.S., but 8 ND, 2 at 

about 100

Sharp and 

Lunder (2004)

20.3 Carpet dust (n = 2) from computer labs in CA CARB (2005)
4-13

Table 4-1. Congener specific concentrations of PBDEs in house dust 

in the United States (units in ng/g dwt) (continued)

Congener Concentration

(ng/g dwt) Comment Reference

5.5, 26.1 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

17.2,15.3, 6.0 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

ND, 293.0 Median, maximum from 11 homes in the Boston, 

MA area

Wu et al. (2007)

1.2, <1.3 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

71 ND Mean 17 homes, Washington, DC Stapleton et al. 

(2005)

ND Carpet dust (n = 2) from computer labs in CA CARB (2005)

75 ND Carpet dust (n = 2) from computer labs in CA CARB (2005)

9.3, 5.3, 3.6 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

77 0.8 Carpet dust (n = 2) from computer labs in CA CARB (2005)

0.1, 0.1 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

1.9, <1.3 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

PentaBDE

85 83 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

100 Mean 10 homes throughout United States, but 8 

ND, 2 at 453 and 544

Sharp and 

Lunder (2004)

51 Carpet dust (n = 2) from computer labs in CA CARB (2005)

28.5, 96.4 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

ND, 787 Median, maximum from 11 homes in the Boston, 

MA area

Wu et al. (2007)
4-14

Table 4-1. Congener specific concentrations of PBDEs in house dust 

in the United States (units in ng/g dwt) (continued)

Congener Concentration

(ng/g dwt) Comment Reference

8.8, 4.7 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

85/155 124.0, 51.8, 

19.2

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

99 1,700 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

ND (400 to 

22,500)

Range of 89 homes sampled in Cape Cod, MA; 

55% detected

Rudel et al. 

(2003)

2,352 Mean 10 homes throughout United States Sharp and 

Lunder (2004)

176, 95 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

880 

(70−3,700)

Median; range from 10 homes in Atlanta Sjodin et al. 

(2008)

776 Carpet dust (n = 2) from computer labs in CA CARB (2005)

612, 2,295 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

2460, 1170, 

536

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

1014, 14979 Median, maximum from 11 homes in the Boston, 

MA area

Wu et al. (2007)

1400, 840 Mean, geometric mean from homes in 

Amarillo/Austin, Texas from 2006 (n = 20)

Harrad et al. 

(2008b)

3830 Median, n = 49 homes in Northern CA Zota et al. (2008)

4430; 

(807−11700)

Median, range, n = 11 homes in Davis, CA Hwang et al. 

(2008)

100 274 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

ND (300) to 

3,400

Range of 89 homes sampled in Cape Cod, MA; 

20% detected

Rudel et al. 

(2003)

911 Mean 10 homes throughout United States Sharp and 

Lunder (2004)
4-15

Table 4-1. Congener specific concentrations of PBDEs in house dust 

in the United States (units in ng/g dwt) (continued)

Congener Concentration

(ng/g dwt) Comment Reference

150 (<8−660) Median; range from 10 homes in Atlanta Sjodin et al. 

(2008)

135 Carpet dust (n = 2) from computer labs in CA CARB (2005)

103, 429 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

436, 204, 77 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

174, 2776 Median, maximum from 11 homes in the Boston, 

MA area

Wu et al. (2007)

240, 160 Mean, geometric mean from homes in 

Amarillo/Austin, Texas from 2006 (n = 20)

Harrad et al. 

(2008b)

756 Median, n = 49 homes in Northern CA Zota et al. (2008)

793; 

165−2,000

Median, range, n = 11 homes in Davis, CA Hwang et al., 

2008

30.6, 16.1 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

116 229 Carpet dust (n = 2) from computer labs in CA CARB (2005)

118 70 Carpet dust (n = 2) from computer labs in CA CARB (2005)

<0.04, <0.04 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

119 12 Carpet dust (n = 2) from computer labs in CA CARB (2005)

126 ND Carpet dust (n = 2) from computer labs in CA CARB (2005)

HexaBDE

138 17 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

181 Mean 10 homes throughout U.S., but 8 ND, 1 at 

1,668

Sharp and 

Lunder (2004)

15 Carpet dust (n = 2) from computer labs in CA CARB (2005)

8, 23 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)
4-16

Table 4-1. Congener specific concentrations of PBDEs in house dust 

in the United States (units in ng/g dwt) (continued)

Congener Concentration

(ng/g dwt) Comment Reference

20.9, 12.1, 5.2 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

ND, 77 Median, ng/g, from 11 homes in the Boston, MA 

area

Wu et al. (2007)

2.28, <0.04 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

153 181 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

243 Mean 10 homes throughout U.S., but 7 ND, 1 at 

1,510

Sharp and 

Lunder (2004)

140 (5−650) Median; range from 10 homes in Atlanta Sjodin et al. 

(2008)

144 Carpet dust (n = 2) from computer labs in CA CARB (2005)

61, 199 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

234.4, 124.2, 

47.0

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

107, 563 Median, maximum from 11 homes in the Boston, 

MA area

Wu et al. (2007)

240, 120 Mean, geometric mean from homes in 

Amarillo/Austin, Texas from 2006 (n = 20)

Harrad et al. 

(2008b)

30.2, 25.9 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

154 156 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

156 Mean 10 homes throughout U.S., but 8 ND, 1 at 

1,050

Sharp and 

Lunder (2004)

80 (<4−260) Median; range from 10 homes in Atlanta Sjodin et al. 

(2008)

95 Carpet dust (n = 2) from computer labs in CA CARB (2005)

54, 189 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005) 
4-17

Table 4-1. Congener specific concentrations of PBDEs in house dust 

in the United States (units in ng/g dwt) (continued)

Congener Concentration

(ng/g dwt) Comment Reference

182.8, 94.4, 

35.0

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

93, 455 Median, ng/g, from 11 homes in the Boston, MA 

area

Wu et al. (2007)

240, 100 Mean, geometric mean from homes in 

Amarillo/Austin, Texas from 2006 (n = 20)

Harrad et al. 

(2008b) 

17.4, 9.6 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

155 8 Carpet dust (n = 2) from computer labs in CA CARB (2005)

156 ND Mean 17 homes, Washington DC Stapleton et al. 

(2005)

166 ND Carpet dust (n = 2) from computer labs in CA CARB (2005)

HeptaBDE

181 ND Carpet dust (n = 2) from computer labs in CA CARB (2005)

183 31 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

60 Mean 10 homes throughout United States, but 9 

ND, 1 at 604

Sharp and 

Lunder (2004)

70 (<4−4,000) Median; range from 10 homes in Atlanta Sjodin et al. 

(2004)

130 Carpet dust (n = 2) from computer labs in CA CARB (2005)

18.6, 19.3 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

27.9, 32.9, 

15.1

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

28, 16 Mean, geometric mean from homes in 

Amarillo/Austin, Texas from 2006 (n = 17)

Harrad et al. 

(2008b)

18, <21 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

184 ND Mean 17 homes, Washington DC Stapleton et al. 

(2005)
4-18

Table 4-1. Congener specific concentrations of PBDEs in house dust 

in the United States (units in ng/g dwt) (continued)

Congener Concentration

(ng/g dwt) Comment Reference

190 5 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

24 Carpet dust (n = 2) from computer labs in CA CARB (2005)

191 ND Mean 17 homes, Washington DC Stapleton et al. 

(2005)

OctaBDE

196 15 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

3.6, 2.6, 3.9 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

197 17 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

2.7, 3.3, 5.6 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

203 109 Carpet dust (n = 2) from computer labs in CA CARB (2005)

3.6, 3.6, 4.9 Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

<3, <3 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

NonaBDE

206 51 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

76.3, 48.1, 

40.5

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

207 30 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

45.9, 25.3, 

26.6

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)
4-19

Table 4-1. Congener specific concentrations of PBDEs in house dust 

in the United States (units in ng/g dwt) (continued)

Congener Concentration

(ng/g dwt) Comment Reference

<0.1, <0.1 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

208 35 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

35.6, 17.5, 

29.4

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

DecaBDE

209 2,090 Mean 17 homes, Washington DC Stapleton et al. 

(2005)

2,394 Mean 10 homes throughout United States Sharp and 

Lunder (2004)

2000 

(120−21,000)

Median; range from 10 homes in Atlanta Sjodin et al. 

(2008)

7500 Carpet dust (n = 2) from computer labs in CA CARB (2005)

665, 8,567 Median, mean (n = 9) from vacuum samples in 

Dallas, TX in 2004

Schecter et al. 

(2005)

4,502, 1,703, 

1,811

Geometric mean of 3 locations/home from 20 

homes in Boston, MA area; living room, bedroom, 

and vacuum

Allen et al. 

(2008)

ND, 9,020 Median, ng/g, from 11 homes in the Boston, MA 

area

Wu et al. (2007)

1,600, 1,300 Mean, geometric mean from homes in 

Amarillo/Austin, Texas from 2006 (n = 17)

Harrad et al. 

(2008b)

2,810, 903 Mean, median from 12 homes in Albany, NY JohnsonRestrepo and 

Kannan (2009)

samples with the highest concentration had the commercial pentaBDE commercial ration of 0.6 

for BDE 47/BDE 99. 

Clothes dryer lint was studied from U.S. and German sources (Schecter et al., 2009b). 

Dryer lint from 12 U.S. and 7 German homes were analyzed for 13 congeners: BDE 17, BDE 28, 

BDE 47, BDE 66, BDE 77, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, 
4-20

BDE 183, and BDE 209. The U.S. homes were in Dallas, and the German homes were in 

Hamburg. The U.S. household lint samples ranged from 321 to 3,073 ng/g dwt, with a median of 

803 and a mean of 1,138 ng/g dwt. The German households ranged from 30 to 3,069 ng/g dwt, 

with a median of 71 and a mean of 361 ng/g dwt. BDE 47, BDE 99, and BDE 209 comprised 

more than 90% of the sum of the 12 congeners in the U.S. samples. For the median, for 

example, BDE 47, BDE 99, and BDE 209 comprised 180, 204, and 273 ng/g dwt of the total of 

803 ng/g dwt. 

 Ten women from a prior Environmental Working Group study on breast milk (of a total 

of 20 women in the earlier study) collected samples of dust from their home (Sharp and Lunder, 

2004). The 10 samples were taken in CA (2 samples), TX, CO, DC, MI, WA, OR, FL, and MT. 

A total of 13 congeners were sampled; detection levels ranged from 50 ng/g dwt for the 

lower-brominated congeners to 90 ng/g dwt for the mid-brominated congeners and finally 

400 ng/g dwt for BDE 209. One of the 10 individuals had very high measurements, at a total of

41,203 ng/g dwt for all congeners; the next highest was 16,366 ng/g dwt. This individual had 

used her vacuum to clean up polyurethane foam residues when she removed carpet padding, two 

mattress pads, and an uncovered foam cushion from her home. This study did not find the 

typical penta BDE commercial blend of 0.6 for BDE 47/BDE 99 in all cases. In half of the 

samples, a significantly higher amount of BDE 47 was found. However, the overall averages of 

the three highest congeners, BDE 47, BDE 99, and BDE 209, were similar to the findings for 

homes in the Washington, DC area: BDE 47 was found at an average of 1,847 ng/g dwt, BDE 99 

at 2,352 ng/g dwt, and BDE 209 at 2,394 ng/g dwt. Concentrations in the dust were not 

correlated to the number of electronic appliances or computers, foam furniture, or recent 

remodeling. 

 BDE concentrations in dust were determined from 10 vacuum bag samples from Atlanta, 

GA, Germany, Australia, and the United Kingdom (Sjodin et al., 2008; first partially reported in 

Sjodin et al., 2004) for a total of 40 samples. These samples were analyzed for BDE 47, 

BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209. The U.S. concentrations were 

uniformly higher than the German and Australian data: the range of total PBDEs (sum of the 

7 congeners) in Germany was 17−550 ng/g dwt (median = 74) and in Australia was 

500−13,000 ng/g dwt (median = 1,200), while in United States, the range was 530−29,000 ng/g 

dwt (median = 4,200). The UK data were the most unique. Concentrations of six congeners, not 
4-21

including BDE 209, were generally low (medians less than 30 ng/g dwt), but the median for 

BDE 209 was 10,000 ng/g dwt with a range of 910−54,000 ng/g dwt. The authors claimed that 

the pentaBDE pattern found in these samples was similar to that in the products; specifically, that 

BDE 99 is similar in concentration to that of BDE 47. This contention is different from those 

made by others, who note that the ratio of BDE 47 to BDE 99 in commercial penta formulations 

is about 0.60. 

Wenning et al. (2006) collected vacuum bag dust and air conditioner filter dust from 

homes in Northern California and Wellington, New Zealand, but the average congener-specific 

concentrations they presented were for all samples, not distinguishing between geography or 

whether it was vacuum or air conditioner-generated dust. The average total BDE concentration 

over 13 samples, including 10 vacuum and 3 air conditioner samples, was 13,570 ng/g dwt, with 

BDE 209 comprising 9,052 ng/g dwt, and BDE 99 and BDE 47 comprising 2,140 and 

1,120 ng/g dwt, respectively. The BDE 209 results were skewed by a business air conditioner 

filter showing 84,500 ng/g dwt, from a total of 96,300 ng/g dwt, so the general 

representativeness of these samples must be questioned. 

Indoor air and dust were sampled in 120 homes in Cape Cod, MA, in two rounds of 

60 samples each starting in 99 and ending in 2001 (Rudel et al., 2003). A total of 89 homes were 

sampled for PBDEs in dust, Only three BDEs (BDE 47, BDE 99, and BDE 100) were analyzed, 

and the reporting limits were high―400 ng/g dwt for BDE 47 and BDE 99 and 300 ng/g dwt for 

BDE 100. Because of this, the data had limited value. BDE 47 was found in about 50% of the 

samples with a maximum of 9,860 ng/g dwt; BDE 99 was found also about 50% of the time with 

a maximum of 22,500 ng/g dwt; and BDE 100 was found 20% of the time with a maximum of 

3,400 ng/g dwt. 

Johnson-Restrepo and Kannan (2009) collected dust samples from the vacuum bags of 

residents of 12 homes in Albany, NY, between December 2007 and January 2008. Twenty 

congeners from BDE 28 to BDE 209 were measured in the dust. BDE 209 was the major 

congener found, accounting for between 26 and 99% of the total concentration. The mean of the 

12 samples was 3,190 ng/g, the median was 1,910 ng/g, and the range was 380 to 9,340 ng/g. 

 Four computer wipe and nine vacuum bag bulk dust samples were taken in Dallas, Texas 

in 2004 (Schecter et al., 2005; date of sampling not given but presumed to be in 2004 as paper 

was submitted in July 2004). The computer wipe samples were in units of ng/100 cm2

, and 
4-22

because no other data were available with which to compare these units, results from this study 

are not provided in Table 4-1. Briefly, concentrations ranged from 77 to 1,536 ng/100 cm2, with 

BDE 209 explaining over 90% in the two samples of PC monitor screens, and about 53% in both 

PC casing samples. In these PC casing samples, BDE 99 and BDE 47 were next in 

predominance, accounting for 25% and 11%, respectively. In nine bulk dust samples, the total 

ranged from 705 to 65,777 ng/g dwt, with a median of 2,507 and a mean of 12,136 ng/g dwt. 

BDE 209 was the prominent congener in seven of nine samples (BDE 99 was the dominant in 

the other two samples), explaining 95% of the concentration in the highest sample 

(65,777 ng/g dwt of 69,283 ng/g dwt total) and 66−87% in the other samples. In one sample, 

however, BDE 209 explained less than 1% of the 30,368 ng/g dwt total, with BDE 99 explaining 

13,841 (46%) and BDE 47 explaining 10,538 ng/g dwt (35%).

 Allen et al. (2008) reports on BDEs in house dust samples that were part of a study 

looking at the correlations between indoor air and bulk dust samples in 20 homes in the urban 

setting of Boston, MA. They took three bulk samples out of each home and developed geometric 

mean congener-specific concentrations for the three common locations: the main living room, 

the bedroom, and a sample from a home vacuum (location undetermined). Like other studies, 

BDE 209 dominated all locations, with the geometric means over these three locations being 

4,702, 1,866, and 1,811 ng/g dwt, respectively. BDE 99 was second highest at 2,460, 1,170, and 

536, respectively, and BDE 47 was the third most prevalent at 1,865, 837, and 338, respectively. 

The geometric means of the total concentrations (including BDE 17, BDE 28/33, BDE 47, 

BDE 49, BDE 66, BDE 75, BDE 85/155, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, 

BDE 183, BDE 196, BDE 197, BDE 203, BDE 206, BDE 207, BDE 208, and BDE 209) of the 

three locations were 13,732, 6,255, and 4,269 ng/g dwt, respectively. They found a correlation 

between air and dust concentrations of the penta formulation congeners (BDE 17 up to 

BDE 154), but did not find a correlation between BDE 209 air and dust concentrations. Their

highest concentration of 544,000 ng/g dwt was dominated by BDE 209, at 527,000 ng/g dwt. 

 In another study of the Boston, MA, area, Wu et al. (2007) sampled 11 houses and 

measured for BDE 17, BDE 28/33, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 138, 

BDE 153, BDE 154, BDE 183, and BDE 209. This was a study in which women’s breast milk 

in these homes were also sampled (results reported in Chapter 5). Results were log normal, with 

a geometric mean of total BDE concentration of 1,910 ng/g dwt. BDE 99 and BDE 47 were 
4-23

quantified in nearly all samples, with median concentrations of 1,010 and 670 ng/g dwt, 

respectively. BDE 209 was quantified in 5 of 11 samples, with quantified concentrations ranging 

between 1,360 and 9,020 ng/g dwt. The median total concentration was 1,910 ng/g dwt. They

found a correlation of r = 0.76 (p = 0.003) between concentrations of pentaBDE congeners in the 

breast milk of first time mothers and dust sampled from their homes. There was also a 

statistically significant correlation between dust levels and with reported dietary habits, 

particularly the consumption of dairy products. 

Harrad et al. (2008b) sampled between 10 and 28 homes in each of these locations: 

Amarillo/Austin in Texas; Birmingham, UK; Toronto, Canada; and Wellington, New Zealand. 

They measured BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and 

BDE 209. The most interesting finding was the profile of BDE congeners found in the different 

locations. The UK site had the highest BDE 209 concentrations, with an average of 45,000, a 

median of 2,800, and a geometric mean of 3,800 ng/g dwt (n = 16), compared to the results from 

Texas, which have an average of 1,600, a median of 1,300, and a geometric mean of 

1,300 ng/g dwt (n = 17). The two highest concentrations of BDE 209 were found in homes in 

the United Kingdom, at 520,000 and 100,000 ng/g dwt (Harrad et al., 2008b). The 

concentrations of the tri-hexa congeners (BDE 28 through BDE 154) were the highest for the 

Texas dust. For example, the geometric mean for the sum of these six congeners from the Texas 

sites was 1,800 compared to 52 from the UK sites. Due to the high BDE 209 findings, the 

overall total concentrations were highest for the United Kingdom, similar to the findings of 

Sjodin et al. (2008). The geometric mean total concentrations for the United Kingdom, United 

States, and Canadian sites were 4,500, 3,600, and 1,200 ng/g dwt. Totals were not provided for 

New Zealand because the BDE 183 and BDE 209 congeners were not measured. Concentrations 

of the tri-hexa congeners totaled 92 ng/g dwt, which was comparable to the tri-hexa totals for the 

United Kingdom. 

Zota et al. (2008) took dust samples in 49 homes in Northern California and measured 

them for BDEs 47, 99, and 100. Median and range of concentrations were (in ng/g dwt): 

BDE 47―3,750 and 11−107,000; BDE 99―3,830 and 102−170,000; and BDE 100―756 and 

<MRL―30,900. As Zota et al. (2008) notes, these concentrations were 4−10 times higher than 

concentrations found for these congeners in U.S. studies summarized here (Rudel et al., 2003; 

Wu et al., 2007; Stapleton et al., 2005; Harrad et al., 2008b) and orders of magnitude higher than 
4-24

studies abroad. The authors also identified blood samples from NHANES 2001/2002 that were 

from California and compared them to the entire NHANES 2001/2002 statistical extrapolations 

for the United States. Their geometric mean for the sum of six congeners (BDE 28, BDE 47, 

BDE 99, BDE 100, BDE 153, and BDE 154) from samples in California was 62 ng/g lipid 

weight (lwt), while the total for all of the United States was 38.6 ng/g lwt. They suggest this 

might have resulted from more stringent standards for flammability in California as compared to 

other states. The NHANES database is discussed further in Chapter 5. 

Hwang et al. (2008) measured seven congeners (BDE 47, BDE 66, BDE 99, BDE 100, 

BDE 138, BDE 153, and BDE 154) in dust collected from vacuum bags provided by residents in 

10 apartments and 1 community hall in Davis, California. They found that the total ranged from 

1,780−25,500 ng/g dwt, with a median of 9,240 ng/g dwt. Median and range of the three 

dominant congeners were (in ng/g dwt) BDE 47―2,920 and 655−8,830; BDE 99―4,430 and 

807−11,700; and BDE 100―793 and 165−2,000. 

The California Air Resources Board (CARB) has conducted a substantial amount of 

sampling for brominated flame retardants in air and dust, both outside near industrial sites and in 

urban and background locations, and inside at workplaces and industrial sites.1

Wilford et al. (2005) studied house dust concentrations from 68 of 74 homes in Canada 

that previously had been studied for air concentrations of PBDEs (see Section 4.5) in the winter 

of 2002/2003. They found that the data were log normally distributed, with a geometric mean of 

2,000 ng/g, a median of 1,800 ng/g, and an arithmetic mean of 5,500 ng/g. For the arithmetic 

mean, BDE 209 comprised 1,000 ng/g of the total, or 18%, while for the median, BDE 209 

comprised half, at 900 ng/g. This suggests that BDE 209 dominates the highest concentrations 

found. After BDE 209 for the mean, the highest concentrations were BDE 99 at 1,800 ng/g and 

BDE 47 at 1,100 ng/g, with other concentrations less than 500 ng/g. 

 Two samples of 

carpet dust in an indoor computer training facility totaled 10,200 and 4,800 ng/g dwt (CARB, 

2005). Congeners in the dust that were not in the air that was sampled in this facility included 

BDE 77, BDE 138, BDE 155, BDE 183, and BDE 203 (BDE 203 and BDE 183 are congeners in 

the octa formulation). The dominant congener was BDE 209, at 7,560 and 2,800 ng/g dwt, 

followed by BDE 99 at 856 and 695 ng/g dwt, and BDE 47 at 502 and 411 ng/g dwt. 

 

1

See http://www.arb.ca.gov/toxics/pbde.htm. 
4-25

 Sampling of house dust has also occurred in foreign countries, including Europe and 

Kuwait. Fabrellas et al. (2005) report sampling done by “Euroconsumers Organization” and 

“CIEMAT-POPs Group,” in which dust was collected from 100 vacuum cleaner bags from Spain 

(34 bags), Belgium (32 bags), Portugal (22 bags), and Italy (12 bags) and measured for BDEs. 

Homologue group concentrations from mono- to heptaBDE were determined; levels of 

detections ranged from 0.006−0.220 ng/g dwt (BDE 3) to 0.55−20.9 ng/g dwt (for BDE 209). 

Results suggest Italy has the highest total concentrations, at 581 ng/g dwt, followed by Portugal 

(354 ng/g dwt), Spain (238 ng/g dwt), and Belgium (190 ng/g dwt). However, it is unknown if 

these mean values are significantly different. The highest value from Spain was from a dust 

sample that had been sucked out from the inside of a computer. Urban house dust samples were 

found to be higher than rural, and BDE 209 was found to comprise greater than 60% of total 

PBDE in all cases. The ratio of BDE 47 to BDE 99, which is 0.5−0.7 in 24% of the samples, is 

the same as the commercial penta formulation, DE-71, which has a ratio of 0.6, suggesting these 

samples were dominated by this commercial mixture. 

Harrad et al. (2006) measured BDEs in house dust at eight homes sampled in 2005 in the 

United Kingdom. They reported an average of 215 ng/g total (maximum of 625 ng/g total), 

which included BDE 28, BDE 47, BDE 49, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, and 

BDE 154 (BDE 209 not measured). The ratio of BDE 47 to BDE 99 was about 0.5. The study 

that Harrad et al. (2008b) later published on different locations around the world (United 

Kingdom, United States, New Zealand, and Canada) included these earlier data. 

Harrad et al. (2008a) later measured dust samples taken from cars, homes, and offices in 

locations in southern United Kingdom (West Midlands area), and found the highest BDE 209 

measurements in dust samples that have been found as of the date of this report, including 

findings of 2,200,000 ng/g in a home and 2,600,000 ng/g dwt in a car (further discussion on the 

car data provided in Section 4.4.2 below). The average total BDE concentration (sum of 

BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209) of 30 dust 

samples taken in homes was 260,000, which was the same concentration provided as the average 

for BDE 209. This was because the concentrations of the other congeners were all less than 

100 ng/g dwt (sum of average congener concentrations was about 150 ng/g dwt), and, hence, 

BDE 209 overwhelmed the average. The average of 18 office samples was 31,000 ng/g dwt 

total, with BDE 209 explaining 30,000 ng/g dwt of that average. They went back to sample one 
4-26

of the two homes with high BDE 209, 1,400,000 ng/g dwt, and found a similar concentration: 

900,000 ng/g dwt 

Karlsson et al. (2007) report on measurements of PBDEs in air, house dust, and blood 

from five households in Sweden. The average total BDE (including 13 BDEs) was about 

690 ng/g dwt, dominated by BDE 209 that was 490 ng/g dwt, followed by BDE 99 and BDE 47, 

at 99 ng/g dwt and 51 ng/g dwt, respectively. Gevao et al. (2005) report on simultaneous air and 

dust from homes in Kuwait in 2005. PUF samplers were used to measure air samples, and bulk 

dust samples were obtained from vacuum bags in 17 homes in Kuwait between 2/29/2004 and 

4/11/2004. Individual congeners measured included BDEs 28, BDE 47, BDE 100, BDE 99, 

BDE 85, BDE 154, BDE 153, and BDE 183 (BDE 209 not measured), although individual 

concentrations were not provided. Total PBDE dust concentrations ranged from 

0.2−24 ng/g dwt (geometric mean 9 ng/g dwt) and were log normally distributed. BDE 47 and 

BDE 99 seemed to track well with what the authors claimed was a commercial penta mixture; 

however, the 47/99 ratio of this mixture was not 0.6 but rather close to 1.0. House dust 

correlated with indoor air―when one was high, the other was also. 

 Recent data from Europe show very high levels of BDE 209. House dust and lint 

samples were collected in Scotland, Northern England, and Germany (Pless-Mulloli et al., 2006) 

and compared with other published data from the United States, although the comparison did not 

include more recent data such as those from Allen et al. (2008). The United Kingdom had the 

highest concentrations, though they were dominated by BDE 209, while Germany’s 

concentrations were much lower. The United Kingdom dust samples had a mean of 11,325 ng/g, 

a median of 3,933 ng/g, and a maximum of 54,858 ng/g. The UK samples were dominated by 

BDE 209, which had an average concentration of 11,233 ng/g (of the 11,325 ng/g total). The 

median of U.S. dust measurements were similar to the median of the United Kingdom’s 

measurements at 4,200 ng/g, but BDE 209 was not nearly as dominant, having a median of only 

2,000 ng/g. As discussed earlier, the data in Allen et al. (2008) from homes in the Boston, MA 

area have higher BDE 209 concentrations than the measurements summarized in Pless-Mulloli et 

al. (2006). 

 In summary, the most common indoor congeners include those associated with the penta 

formulations, BDE 47 and BDE 99, and the single congener most associated with the deca 

formulation, BDE 209. Concentrations of total BDEs ranged from the low 100s to well over 
4-27

10,000 ng/g dwt. Concentrations of BDE 209 have been found in the hundreds of thousands to 

over two million of ng/g dwt in the United Kingdom (Harrad et al., 2008a) and in the hundreds 

of thousands in one study in the U.S. (from Allen et al., 2008). Most authors reported that the 

results suggested the presence of the penta formulation, based on the ratio of the two congeners, 

BDE 47 and BDE 99. However, there was not uniform agreement on the ratio: some claimed 

that comparable amounts of the two congeners suggest the presence of the commercial product, 

while others claim that the penta formulation translates to a consistent 47/99 ratio of about 0.6. 

When quantified, BDE 209 dominated the profile above BDE 47 and BDE 99, with 

concentrations generally in the 2,000−10,000-ng/g dwt range, although with some outlier 

concentrations above 50,000 ng/g dwt, and as noted, some well above that in the hundreds of 

thousands to millions of ng/g dwt. Dust concentrations of BDEs appear much higher in 

U.S. dust, well into the range of thousands of ng/g dwt for total concentrations, compared with 

limited measurements in Europe, which appear to be limited to the hundreds of ng/g dwt. 

Exceptions to this generalization include data from the United Kingdom showing substantial 

concentrations of BDE 209, and one study from Kuwait, which had a high concentration of total 

BDEs just above 100 ng/g dwt. 

4.4.2. Car and Airplane Dust Concentrations

 High concentrations of BDE congeners in cars and airplanes recently have been found. 

The authors of the studies have noted that the concentrations found are higher than those 

typically found in homes. This becomes important even though the time spent in these vehicles 

is small relative to the time spent in a home. 

The Harrad et al. (2008a) West Midland, United Kingdom study mentioned in the 

previous section found the highest BDE 209 measurements in dust samples that have been found 

as of the date of this report, including a finding of 2,600,000 ng/g dwt in a car. The average 

concentration for 20 car samples was 410,000 ng/g dwt (std. deviation 770,000; personal 

communication with Harrad). Cars had statistically significantly higher concentrations of lower 

BDE congeners, tri-hexaBDEs, in the hundreds of ng/g for BDE 47, BDE 99, BDE 100, 

BDE 153, and BDE 154, as compared to offices and homes. They went back to sample the high 

car finding of BDE 209, 2,600,000 ng/g dwt, and found a similar concentration: 2,200,000 ng/g 

dwt. 
4-28

 Lagalante et al. (2009) sampled dust within 60 cars in the United States and, like Harrad 

et al. (2008a), found a dominance of BDE 209. The cars were manufactured mostly in the 

2000s; one car each was manufactured in 1990, 1998, and 1999. The median BDE 209 level was 

48,100 ng/g dwt, and it comprised 95% of total BDEs measured. Median concentrations of the 

other congeners measured were BDE 28―118 ng/g dwt, BDE 47―880 ng/g dwt, 

BDE 99―1,130 ng/g dwt, BDE 100―211 ng/g dwt, BDE 153―163 ng/g dwt, and 

BDE 183―73 ng/g dwt. All of these cars were from car dealerships and were cleaned. 

Significant relationships between concentrations and car descriptors were found for BDE 209 

and these descriptors: vehicle model year, vehicle manufacturer, and country of origin; there was 

no relationship by dealership, suggesting that cleaning procedures might not influence BDE 

levels in the cars. Even though there was significant differences between concentrations by 

manufacturers, up to 3 orders of magnitude difference were seen within some of the 

manufacturers, suggesting that there were variables explaining the concentrations not 

determined, or else there simply was variability within the cars. 

Christiansson et al. (2008) studied levels of dust in planes making intercontinental flights 

out of Sweden to Japan, China, Poland, Canada, United States, United Kingdom, and to India. 

Nine individuals provided the dust samples on their intercontinental trips, collecting the dust 

mostly from ventilation ducts in the airplane restrooms (no further detail provided). These nine 

individuals also provided blood samples before and after their trips to determine whether 

exposures may have resulted in an elevation of PBDEs. Two additional individuals provided 

dust samples from one of their frequent trips within Sweden, and one blood sample each. All the 

results in this study were provided in units of pmol/g dwt for dust and pmol/g lwt for blood. To 

convert to the units used in this report, the results were multiplied by the molecular weight and 

0.001 (to convert pmol to pg, multiply pmol × pg/pmol, or essentially, g/mol, which is the 

molecular weight; and then from pg to ng, multiply by 0.001). The median and maximum of the 

congeners in dust were (in ng/g dwt) BDE 28―140 and 2,800; BDE 47―3,500 and 228,000; 

BDE 99―3,800 and 291,000; BDE 100―900 and 179,000; BDE 153―502 and 23,000; 

BDE 154―370 and 63,000; BDE 183―490 and 193,000; and BDE 209―17,000 and 190,000. 

The authors display a table and discuss how these concentrations are higher than those found in 

house dust from other countries, including the United States, citing three U.S. studies (all 

discussed above). They note that the travelers showed blood levels similar to Swedish general 
4-29

population, but that levels of most of the congeners increased upon return from their 

intercontinental trip. BDE 209 had the highest mean and median concentration of all congeners, 

at 3.4 and 2.6 ng/g lwt, respectively. The second highest was BDE 47 at 1.8 and 1.2 ng/g lwt, 

respectively. As discussed in Chapter 5, NHANES pooled blood samples showed a level of 

2.0 ng/g lwt of BDE 209, but that the median for individual samples for BDE 47 was 20.5 ng/g 

lwt. 

 

4.5. AIR CONCENTRATIONS

Efforts to monitor PBDEs in outdoor air have been more extensive than efforts to monitor 

indoor air in the United States, despite evidence that shows indoor air to have generally higher 

concentrations of BDEs than outdoor air. Outdoor measurements have been taken in California, 

in the Great Lakes Region, and in several other states including New York, Louisiana, Indiana, 

Maryland, and others. Only one study focusing on indoor air in residential settings, in the 

Boston, MA area, could be found. Sites in Europe and Asia suggest lower concentrations as 

compared to the United States, although one study from China reports much higher 

concentrations. One study reported on air concentrations within cars in Greece. Section 4.5.1 

describes efforts to monitor outdoor air, while Section 4.5.2 describes indoor air studies. 

Table 4-2 provides congener-specific concentrations for both outdoor and indoor air in the 

United States. 

4.5.1. Outdoor Air Concentrations

CARB established the California Ambient Dioxin Air Monitoring Network. This 

Network, composed of 11 sites, began sampling in 2002, and the data for the years 2002−2004 

are posted at http://www.arb.ca.gov/pub/dioxin/cadamp.php. PBDE data were taken from 7 of 

these 11 sites, 4 from the Bay Area and 3 from the South Coast, beginning in 2003. There were 

6 monthly samples in 2003 and 12 monthly samples in 2004. Individual site data and 
4-30

Table 4-2. Outdoor and indoor congener-specific air concentrations of 

PBDEs in the United States (units in pg/m3

) 

Congener Concentration

(pg/m3) Comment Citation

TriBDE

17 3.6 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

46 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

1.5−5.6 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

35−106 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

0.3−5.8 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

1.7 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

7.6, 8.1, 7.0 Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal”, bedroom, and living room

Allen et al. 

(2007)

25 2.3 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

29 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

0.4−4.7 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

16−103 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

0.5−3.1 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

28 99 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

2.9−17.8 Range, n = 12 (3 days, 4 samplers) outdoors surrounding

electronics recycling facility in CA

CARB (2005)

102−372 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

0.4−7.4 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

3.0 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)
4-31

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

80, 50 Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

28/33 4.0 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

29.6, 27.3, 25.4 Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal”, bedroom, and living room

Allen et al. 

(2007)

30 0.2 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND−0.9 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

32 ND Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND−0.9 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

33 ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND−1.4 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

0.7−35 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)
4-32

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

35 0.4 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND−1.1 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND−18.1 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

TetraBDE

37 2.3 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

1.3−8.4 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND−468 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

0.6−2.4 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

47 6.2 Michigan (n = 35), 2002/2003 Hoh and Hites 

(2005)

17.4 Chicago (n = 28), 2002/2003 Hoh and Hites 

(2005)

7.0 Indiana (n = 38), 2002/2003 Hoh and Hites 

(2005)

9.2 Arkansas (n = 30), 2002/2003 Hoh and Hites 

(2005)

6.9 Louisiana (n = 26), 2002/2003 Hoh and Hites 

(2005)

5.0 3 Rural/remote sites in MI, NY (n = 36), 1997−1999 Strandberg et al. 

(2001)

33 Chicago, n = 12, 1997−1999 Strandberg et al. 

(2001)
4-33

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

175 (671) Lewes, Del Marva, MD (n = 95) geom. mean (max), vapor 

phase only

Goel et al. 

(2006)

9.7 (26) Horn Point, Del Marva, MD (n = 98) geom. mean (max), 

vapor phase only

Goel et al. 

(2006)

17 (52) Dover, Del Marva, MD (n = 47) geom. mean (max), vapor 

phase only

Goel et al. 

(2006)

34.5 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

1,065 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

30−128 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

604−2,850 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA; average = 1,772

CARB (2005)

30−88 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

53.0 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

226.8, 157.9, 

145.1

Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal”, bedroom, and living room

Allen et al. 

(2007)

740, 420 Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

49 1.2 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

59 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

4.6−36.7 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

449−2,860 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA; average = 1,764

CARB (2005)

2.7−10.5 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

49 9.1, 6.0, 7.2 Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal”, bedroom, and living room

Allen et al. 

(2007)

65 2.0 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)
4-34

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

66 1.6 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

29 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

2.9−10 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

103−750 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

2.3−5.1 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

3.7, 3.5, 3.5 Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal”, bedroom, and living room

Allen et al. 

(2007)

110, 90 Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

71 1.7 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

75 0.4 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

6.6 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND−194 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

0.5−2.4 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)
4-35

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

77 ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

22.1−75.5 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND−2.4 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

2.0 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

ND, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

PentaBDE

85 1.1 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

13 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND−6 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

87−284 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

2.1−9.0 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

2.3 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

3.8, 2.7, 2.5 Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal”, bedroom, and living room

Allen et al. 

(2007)

ND, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

99 5.1 Michigan (n = 35), 2002/2003 Hoh and Hites 

(2005)

7.4 Chicago (n = 28), 2002/2003 Hoh and Hites 

(2005)

5.1 Indiana (n = 38), 2002/2003 Hoh and Hites 

(2005)
4-36

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

5.4 Arkansas (n = 30), 2002/2003 Hoh and Hites 

(2005)

3.0 Louisiana (n = 26), 2002/2003 Hoh and Hites 

(2005)

3.4 3 Rural/remote sites in MI, NY (n = 36), 1997−1999 Strandberg et al. 

(2001)

16 Chicago, n = 12, 1997−1999 Strandberg et al. 

(2001)

26 (178) Lewes, Del Marva, MD (n = 95) geometric mean (maximum), 

vapor phase only

Goel et al. 

(2006)

5.3 (26) Horn Point, Del Marva, MD (n = 98) geometric mean 

(maximum), vapor phase only

Goel et al. 

(2006)

7.7 (17) Dover, Del Marva, MD (n = 47) geometric mean (maximum), 

vapor phase only

Goel et al. 

(2006)

12.5 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

239 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

9−85 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

402−3,210 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA; avg = 1,771

CARB (2005)

21−111 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

51.0 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

110.8, 66.9, 60.3 Geometric mean of 20 homes, 3 locations/home in Boston,

MA area; “personal”, bedroom, and living room

Allen et al., 

2007

210, 160 Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

100 1.1 Michigan (n = 35), 2002/2003 Hoh and Hites 

(2005)

1.8 Chicago (n = 28), 2002/2003 Hoh and Hites 

(2005)

1.0 Indiana (n = 38), 2002/2003 Hoh and Hites 

(2005)
4-37

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

1.1 Arkansas (n = 30), 2002/2003 Hoh and Hites 

(2005)

0.7 Louisiana (n = 26), 2002/2003 Hoh and Hites 

(2005)

0.5 3 Rural/remote sites in MI, NY (n = 36), 1997−1999 Strandberg et al. 

(2001)

2.0 Chicago, n = 12, 1997−1999 Strandberg et al. 

(2001)

17 (73) Lewes, Del Marva, MD (n = 95) geometric mean (maximum), 

vapor phase only

Goel et al. 

(2006)

5.4 (5.4) Horn Point, Del Marva, MD (n = 98) geometric mean 

(maximum), vapor phase only

Goel et al. 

(2006)

5.3 (5.3) Dover, Del Marva, MD (n = 47) geometric mean (maximum), 

vapor phase only

Goel et al. 

(2006)

4.8 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

100 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

3−17 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

69−448 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

5.2−21.5 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

13.0 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

22.2, 14.4, 12.0 Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal,” bedroom, and living room

Allen et al. 

(2007)

60, 50 Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

116 0.2 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

14.7 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)
4-38

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

ND Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

118 0.3 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND−5.3 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

67−343 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

1.1−3.9 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

ND, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

119 1.4 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

126 ND Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)
4-39

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

ND Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

HexaBDE

138 0.9 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND−6.8 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

95−346 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

1.3−3.9 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

ND, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

153 0.20 3 Rural/remote sites in MI, NY (n = 36), 1997−1999 Strandberg et al. 

(2001)

0.53 Chicago, n = 12, 1997−1999 Strandberg et al. 

(2001)

2.0 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

11 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

3−150 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

1,120−8,900 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA; average = 5,623

CARB (2005)

11.3−33.2 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

3.9 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

8.6, 4.0, 3.5 Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal”, bedroom, and living room

Allen et al. 

(2007)

60, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)
4-40

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

154 0.12 3 Rural/remote sites in MI, NY (n = 36), 1997−1999 Strandberg et al. 

(2001)

041 Chicago, n = 12, 1997−1999 Strandberg et al. 

(2001)

ND For Lewes, Horn Point, Del Marva, MD (n = 240), vapor 

phase only

Goel et al. 

(2006)

2.8 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

13 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

1.5−86.7 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

1,230−5,260 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA; average = 3,455

CARB (2005)

3.8−36.9 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

4.0 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

9.1, 6.1, 5.2 Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal,” bedroom, and living room

Allen et al. 

(2007)

ND, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

155 ND Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND−0.6 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

166 ND Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)
4-41

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND−2.5 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

35−463 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

HeptaBDE

181 ND Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

ND Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

183 1.4 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

6−456 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

5,610−36,700 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA; average = 23,813

CARB (2005)

4.0−32.6 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

1.4 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

ND, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

190 <0.06 (DL) 3 Rural/remote sites in MI, NY (n = 36), 1997−1999 Strandberg et al. 

(2001)
4-42

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

<0.06 (DL) Chicago, n = 12, 1997−1999 Strandberg et al. 

(2001)

ND Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND−16 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA

CARB (2005)

288−1,300 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA

CARB (2005)

ND−3.8 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA

CARB (2005)

OctaBDE

203 ND Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

ND Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

ND, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

NonaBDE

207 ND, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

DecaBDE

209 1.4 Michigan (n = 35), 2002/2003 Hoh and Hites 

(2005)

60.1 Chicago (n = 28), 2002/2003 Hoh and Hites 

(2005)

2.2 Indiana (n = 38), 2002/2003 Hoh and Hites 

(2005)

9.0 Arkansas (n = 30), 2002/2003 Hoh and Hites 

(2005)

2.6 Louisiana (n = 26), 2002/2003 Hoh and Hites 

(2005)
4-43

Table 4-2. Outdoor and indoor congener specific air concentrations of 

PBDEs in the United States (units in pg/m3) (continued)

Congener Concentration

(pg/m3) Comment Citation

<0.10 (DL) 3 Rural/remote sites in MI, NY (n = 36), 1997−1999 Strandberg et al. 

(2001)

0.30 Chicago, n = 12, 1997−1999 Strandberg et al. 

(2001)

10.6 Outdoors at UC Davis, 2004; average of 2 days of filter/PUF 

and filter/XAD samplers

CARB (2005)

58 Indoors in computer facility, average of 2 days, 

8 measurements, 6 with computers on, 2 with computers off, 

in CA

CARB (2005)

140−11,400 Range, n = 12 (3 days, 4 samplers) outdoors surrounding 

electronics recycling facility in CA; mean = 2,764

CARB (2005)

79,700−833,000 Range, n = 6 (3 days, 2 samplers) indoors in electronics 

recycling facility in CA; average = 423

CARB (2005)

123−1,940 Range, n = 9 (3 days, 3 samplers) outdoors in auto shredder 

facility in CA; average = 2,403

CARB (2005)

25.0 n = 84 (12 sampling dates, 7 monitors); average for 2004 in 

Bay Area and South Coast of CA

CADAMP 

(2006)

174, 95, 94 Geometric mean of 20 homes, 3 locations/home in Boston, 

MA area; “personal,” bedroom, and living room

Allen et al. 

(2007)

ND, ND Mean, median indoor concentrations from 12 homes in 

Albany, NY

JohnsonRestrepo and 

Kannan (2009)

site/statewide averages are available from the Web site. Twelve congeners were measured, 

including BDE 17, BDE 28, BDE 47, BDE 65, BDE 77, BDE 85, BDE 99, BDE 100, BDE 153, 

BDE 154, BDE 183, and BDE 209. BDE 47 and BDE 99 had similar 12-month averages in 

2004, of 53 and 51 pg/m3

, respectively, with BDE 209 coming in third at 25 pg/m3. BDE 100 

averaged 13 pg/m3, and others ranged from 0.04 to 4.0 pg/m3. The total average total 

concentration for 2004 was 160 pg/m3. 

 CARB (2005) also sponsored a research monitoring effort that included indoor and 

outdoor sampling at industrial sites, office sites, and outdoors at the University of California at 

Davis (UC Davis). This overall research program was also published in the open literature in 

Cahill et al. (2007). The UC Davis sampling was conducted to evaluate the effectiveness of two 

different active outdoor samplers: one included a filter (for capture of particle bound PBDEs), 
4-44

followed by a PUF (for vapor phase PBDEs), and the second sampler included a filter followed 

by an XAD-2 resin. Four sampling units, two of each, took two samples per day on 

2 consecutive days, 3/17/2004 and 3/18/2004, providing a total of eight samples. Statistical 

analysis suggested the two measured comparably, except that the filter/PUF setup could not 

measure BDE 209 due to analytical difficulties. A total of 33 congeners were measured, making 

it the most substantial air study found in the literature. Total BDE concentration outdoors in this 

test averaged 38 pg/m3 for the filter/PUF setup and 93 pg/m3 for the filter/XAD setup, where 

BDE 209 made up 10 pg/m3 of that total. BDE 47 and BDE 99 dominated the profiles, 

averaging 34.6 and 12.5 pg/m3, respectively, between the two sampling units. 

 Hoh and Hites (2005) took air samples every 12 days at five locations from Lake 

Michigan to the Gulf of Mexico, between August 2002 and December 2003, covering most (but 

not all) months of the year. One site was urban (Chicago), two sites were in remote locations in 

Michigan and Louisiana, one site was in an agricultural region, and one was in the university city 

of Bloomington, Indiana. Although numerous congener data were evaluated, individual data 

were only available for BDE 47, BDE 99, BDE 100, and BDE 209 (these four contributed about 

80% of the total PBDE). The highest congener individually was BDE 47, averaging between 

7 and 17 pg/m3

 over all sites, while the other congeners mostly averaged under 7 pg/m3. The key 

exception was BDE 209, which was important in Chicago, averaging 60 pg/m3 while it was 

9 pg/m3 or less in other sites. Overall, the total PBDE concentration at the Chicago site was the 

highest, averaging 100 pg/m3

; the concentrations at the other four sites were comparable, 

averaging under 30 pg/m3. 

Strandberg et al. (2001) monitored BDEs at four sampling sites including downtown 

Chicago (urban site), Sleeping Bear Dunes, MI (rural site), Sturgeon Point, NY (rural), and Eagle 

Harbor, MI (remote). Four samples per year (taken May through October) for each of the 

four sites and for 3 years (1997, 1998, and 1999) resulted in a total of 48 samples. Congeners 

included BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 190, and BDE 209. The 

average total BDE for Chicago was about 50 pg/m3 over 3 years, while it was about 5−15 pg/m3

for the other three sites. BDE 209 was not detected at the three rural/remote sites but detected at 

levels only at 0.3 pg/m3

 in Chicago. This is in stark contrast to the monitoring of Hoh and Hites

(2005), which showed an average of 60 pg BDE 209/m3

 in 2002/2003. This could likely be due 

to the fact that Strandberg et al. (2001) did their measurements between 1997 and 1999, prior to 
4-45

the time when the penta and octa formulations were taken off the U.S. market, and the market 

became dominated by the deca formulation (which contains 97% BDE 209). Also, Strandberg et 

al. (2001) note analytical difficulties with BDE 209 (degradation during analysis, low recovery) 

and discusses other possibilities for low findings in air (low mobility from source, etc.); Hoh and 

Hites (2005) note much higher BDE 209 in their data and attribute it to greater usage of 

decaBDE, not analytical problems. Gas/particle partitioning suggests 80% of BDE 47 was in the 

vapor phase, 55−65% in the vapor phase for BDE 99 and BDE 100, and only about 30% in vapor 

phase for BDE 153 and BDE 154. 

The sampling network reported on by Strandberg et al. (2001) and Hoh and Hites (2005) 

continued to be monitored during 2005 and 2006, and PBDE results from that effort are reported 

by Venier and Hites (2008). The total PBDE concentrations were highest at the “urban” sites of 

Chicago and Cleveland, with mean concentrations of 65 and 87 pg/m3, respectively, with the 

higher concentrations at Cleveland due to the presence of higher concentrations of BDE 209 in 

several samples. The mean concentrations at the “rural” sites of Sturgeon Point and Sleeping 

Bear Dunes were 9.2 and 8.1 pg/m3, respectively, and the mean concentration at the “remote” 

site of Eagle Harbor was 5.8 pg/m3. The authors also studied the time trends of BDE 47, 

BDE 99, and BDE 209 from data starting in 2003 and observed that BDE 47 and BDE 99 were 

declining rapidly, but that BDE 209 was not declining at any of the five sites. 

Three sites within the Chesapeake Bay area were sampled for BDEs as part of a larger 

effort at air quality monitoring in the region (Goel et al., 2006). A total of 240 samples were 

taken between 2001 and 2003, using high volume samplers to measure both particle and vapor 

phases for BDE 47, BDE 99, BDE 100, and BDE 154. BDEs were detected in 75% of the 

samples, but detection frequency was highest at Lewes, DE 98%, compared to Dover, DE 77%, 

and Horn Point, MD 52%. Vapor phase concentrations were presented and were highest at 

Lewes, with geometric mean concentrations over the 3-year period of 174 pg/m3

, followed by 

Dover and Horn Point at 19 and 10 pg/m3, respectively. It is not clear why particle phase 

concentrations were not presented in Goel et al. (2006), except that they were much less 

frequently found (<5% of samples from Horn Point and Dover), but at 40−50% of samples at 

Lewes. BDE 47 was the highest found in all three sites, from about twice as high as BDE 99 and 

BDE 100, in Horn Point and Dover (10−20 pg/m3 for BDE 47 compared to 5−8 pg/m3 for 

BDE 99 and BDE 100 as geometric means), to about 7−10 times as high at Lewes (175 pg/m3 for 
4-46

BDE 47 compared to 17−26 pg/m3 for BDE 99 and BDE 100). BDE 154 was essentially not 

detected in either vapor or particle phases. The finding at Lewes was attributed to use of spray 

irrigation of municipal wastewater near the sampling site. 

 One of the CARB sites involved outdoor sampling near a possible source of PBDE 

release―near an autoshredder. On this site, there were four samplers, including two upwind and 

two downwind, sampling for 3 days, resulting in 12 samples. There were no field activities 

during the first day of sampling, and the results were lower that day when compared to days 

afterward―when the facility was operating. Another key observation was that samples were 

much higher downwind compared to upwind. However, even the upwind sites were higher than 

the UC Davis site discussed earlier, suggesting that there was offgassing from previously 

deposited BDEs from the nearby source. The highest measurement was for BDE 209, averaging 

over 2,400 pg/m3 for four sampling dates. Only one other sample had at least one reading above 

100 pg/m3, and that was for BDE 99, which had the second highest readings overall, with an 

average of 165 pg/m3 over the 9 days. BDE 47 had the next highest, averaging 64 pg/m3. 

 A second CARB site was near another source: an electronics recycling plant. Three days 

of outdoor sampling with 4 samplers (2 in front and 2 in back) for a total of 12 samples, 

measured the impact of the release of PBDEs from electronics within this recycling facility. 

Samplers located in front, near the loading dock, which was open most of the time, had higher 

concentrations as compared to the back two samplers. BDE 209 was the highest congener 

measured, averaging 2,764 pg/m3

, with the next highest BDE 183 averaging 116 pg/m3, and then 

BDE 47 averaging 70 pg/m3. The interesting thing about BDE 209, which exists predominantly 

in the particle form, is that the concentrations were 40-fold higher in the front samplers, near the 

open doors to the indoor recycling facility. This suggests that near a source (such as near the 

front doors in a recycling facility where BDE 209 could be released), BDE 209 concentrations 

can be extremely elevated, but far from a source (in the back of the building), concentrations are 

much lower. 

 Another study evaluating outdoor air impacts near a source occurred in the United 

Kingdom. Air concentrations before and after a major bonfire were analyzed for both PAHs and 

PBDEs in November of 2000 (Farrar et al., 2004). Guy Fawkes Day (also referred to as “Bonfire 

Festival” in the article), which is centered on the use of fireworks and the lighting of bonfires, 

occurs every year on November 5. Three samplers were set up in the garden of a home in a 
4-47

residential area. They were not in the immediate vicinity of any public bonfires, but there were 

bonfires known to be occurring in the general area. Daily samples were taken from November 1 

to 13, and the evidence clearly showed a rise in concentration on the 5th, with background levels 

on the 4th and before and from the 6th onward. A total of 21 BDE congeners were measured, 

although BDE 209 was not measured. Background concentrations were only about 4 pg/m3, with 

quantifiable measurements of BDE 47 (2 pg/m3

), BDE 99 (1.5 pg/m3), and BDE 100 (0.5 pg/m3). 

This clean air was attributed to air originating over the Atlantic Ocean. Concentrations reached 

95 pg/m3 on November 5, with the highest concentrations of BDE 99 (14 pg/m3), BDE 153 

(13 pg/m3), BDE 154 (10 pg/m3), BDE 166 (10 pg/m3), BDE 47 (8 pg/m3), BDE 49 (8 pg/m3), 

and similar concentrations of 4 pg/m3 for BDE 66, BDE 85, BDE 100, BDE 181, and BDE 190. 

The authors suggest that likely sources of the PBDEs were the burning of discarded clothing or 

furniture, with temperatures not hot enough to destroy the PBDEs. No explanations were 

provided for the predominance of BDE 153 and BDE 154, which is different from the more 

typical predominance of BDE 47 and BDE 99 (as seen in background air). 

 Gouin et al. (2006) studied particle bound air transport of BDE 209 in Southern Ontario 

by air sampling with a PS-1 sampler during the period between 1/23/2002 and 6/5/2002. While 

the focus was on BDE 209, BDE 47, BDE 99, and BDE 100 were also measured in 28 samples

taken roughly weekly during this time frame. Concentrations of BDE 209, exclusively found in 

the particle phase, ranged from nondetect (detection limit = 3 pg/m3

) to 105 pg/m3, with a mean 

of 19 pg/m3

. BDE 99 and BDE 100 were rarely detected, while BDE 47 was detected in nearly 

every sample with concentrations of mostly less than 5 pg/m3, dominated by the vapor phase (as 

measured in the PUF portion of the sampler) in the warmer time frames, but with higher 

concentrations in the particle phase (as measured by the amounts on the glass fiber filter) from 

January to about the end of March. Blanchard et al. (2004) also reported on the concentrations 

of BDE 209 in Canadian air. They found air concentrations averaging 1.8 pg/m3 of BDE 209 

(predominantly vapor phase) for 72-hour samples collected every month during the year. 

Wilford et al. (2008) sampled air at a semirural site in northwest England in April to May 

of 2004. Their analysis focused almost exclusively on characterizing the particle phase of BDEs, 

with a particular emphasis on BDE 209 and other higher-brominated congeners. They did 

measure for the vapor phase using a PUF sampler for a 7-day period and found very low levels 

of BDE 209 in a limited number of samples. More importantly, they stated that the vapor phase 
4-48

was dominated by tri- to hexa BDE congeners, and that these congeners were found at vapor 

phase concentrations at least 2 orders of magnitude higher than in the particle phase. However, 

they did not provide any results for these vapor phase concentrations, or for total (vapor

+ particle) phase concentrations. Their mean and median concentrations of total BDEs in the 

particle phase over 28 samples were 41 and 18 pg/m3

, respectively. The mean and median 

BDE 209 concentrations were 20 and 13 pg/m3

, respectively, showing the dominance of this 

congener in the particle phase profile. In fact, the sum of hepta thru deca congeners, or more 

specifically, the sum of BDE 183, BDE 196, BDE 197, BDE 206, BDE 207, BDE 208, and 

BDE 209, explained over 90% of the total profile over all samples. Concentrations of the 

lower-brominated congeners in the particle phase were mostly less than 1 pg/m3. BDE 183 was 

found in 10 of the 28 samples, at a mean of 4.6 pg/m3

, at a high of 92 pg/m3. 

 Harrad and Hunter (2006) used PUF disk passive samplers to measure the concentration 

of six BDE congeners―BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154―in the 

West Midlands area of the United Kingdom, which includes the city of Birmingham and 

neighboring rural locations. They took air samples at 10 sites over the course of 1 year between 

2003 and 2004. The found average total concentrations (sum of the 6 congeners) at these 10 

sites to range from 2.84 to 23.3 pg/m3, with the highest concentration associated with the city of 

Birmingham. They found ratios of BDE 47 to BDE 99 to range narrowly from 2.95 to 3.57. 

They suggested that by using PUF samplers, they might be capturing mostly the vapor phase of 

the BDE congeners in air. 

Jaward et al. (2004, 2005) have provided a comprehensive sampling of outdoor air in 

both Asia (Jaward et al., 2005) and Europe (Jaward et al., 2004). They used PUF disk samplers 

in a passive mode, meaning they leave the PUF exposed to outdoor air to capture PBDEs for a 

period of 6 weeks, then analyzed the PUFs. Two drawbacks from this approach are that it 

purports to capture only the vapor phase BDEs (it was stated that the PUF captures some 

particles, but the amount of particles was hard to ascertain), and, secondly, that BDE 209 was not 

analyzed in their studies. In their Asian study, Jaward et al. (2005) employed PUF samplers 

simultaneously at 77 sites and measured the air between 9/21/2004 and 11/16/2004. Rural and 

urban sites were sampled in China (32 samples), Japan (20), South Korea (15), and Singapore 

(10). Congeners measured include BDE 17, BDE 28, BDE 32, BDE 47, BDE 49, BDE 75, 

BDE 99, and BDE 100. Total BDEs ranged from 0.1 to tens of pg/m3

, which the authors note is 
4-49

consistent with measurements for the remote coast of Ireland and rural/semirural England. 

However, some samples in China measured up to 340 pg/m3 in an industrial city known to 

manufacture electronics, which is similar to values reported in urban United Kingdom, where 

PBDE usage is known to be high. Mean or median concentrations were not provided. Only 

ranges were given and, as such, trends between congeners cannot be described. With a DL of 

0.13 pg/m3, maximums found include (in units of pg/m3) the following: BDE 17: 35 in China, 

<1.7 otherwise; BDE 28: 130 in China, 6, 52, and 2.6 otherwise; BDE 32: 13 in China, <1.2 

otherwise; BDE 47: 78 in China, <10 otherwise; BDE 49: 48 in China, <3 otherwise; BDE 75: 

13, 12, 19, 1.2 in all four locations; BDE 99: 50 in China, <10 otherwise; BDE 100: 5.5 in China, 

<2.3 otherwise. It was stated that BDE 47 and BDE 99 dominated the profile, contributing 

around 75% to the overall burden. A similar approach was taken in Europe: 71 PUF samplers 

were deployed in 22 European countries, including 25 in urban locations and 46 in rural/remote 

locations. BDEs measured and the results, percent detected in parentheses, and ranges in pg/m3

are as follows: BDE 28 (82%): <0.5−30; BDE 47 (55%): <8−80; BDE 49 (30): <0.5−12; 

BDE 75 (54%): <0.5−3; BDE 99 (45%): <10−120; BDE 100 (41%), <2−20; BDE 153 (55%): 

<0.7−15; and BDE 154 (44%): <0.8−10. Generally, there was over a 700-fold ratio between 

high and low total measurements, with the highest being in urban locations in the United 

Kingdom and the lowest in remote areas of Iceland, Ireland, Norway, and Sweden. Generally, 

low levels were found in Eastern Europe. As in Asia, BDE 47 and BDE 99 constituted about 

75% of the total PBDEs. 

 A study measuring BDEs in China suggested much higher concentrations compared to 

the measurements of Jaward et al. (2005) discussed above. Chen et al. (2006) reported on results 

from 32 pairs of samples collected from four sites―2 industrial, 1 urban, 1 background―in 

China during 6/15/2004−6/30/2004. High volume polyurethane foam/glass fiber filter PS-1 

samplers were used, and detection limits were 0.14−0.58 pg/m3 for BDE 28, BDE 47, BDE 66, 

BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, and BDE 183, and 14.3 pg/m3 for 

BDE 209. BDE 209 dominated the profile, averaging 4,200, 750, 264, and 478 pg/m3 for the 

two industrial, urban, and background settings while the sum of the other PBDEs 

correspondingly averaged 3,673, 230, 89, and 105 pg/m3

. The results appeared to be correlated 

to wind patterns. When blowing in from the west, where there were several big electronic 

markets (where computers are assembled and dismantled), BDE 209 (which is used in computer 
4-50

circuitry and plastic housing) was very dominant; whereas when wind blew in from the 

southeast, directly from industrial areas, the sum of the other BDEs dominated. Congener 

patterns in one of the industrial sites, the urban sites, and the background sites suggests nearby 

use of the commercial formulation, with the high readings of BDE 209, and the nearby use of the 

commercial penta formulation, with BDE 47 and BDE 99 comprising greater than 50% of the 

total. 

 In summary, the major trends seen from the air monitoring studies include the following: 

(1) Outdoor concentrations measured in the United States tended to be in the range of 

20−200 pg/m3 for total BDEs. (2) The profile was dominated by BDE 47 and BDE 99, 

suggesting a penta formulation influence; however, when BDE 209 was measured, it was seen to 

have concentrations equal or greater than BDE 47 and BDE 99. One study in Chicago had 

BDE 209 at concentrations six times higher than BDE 47 or BDE 99, and a study in China also 

showed a dominance of BDE 209 in air profiles of industrial, urban, and background settings. 

(3) Concentrations were found to be higher in industrial/urban settings as compared to 

rural/background settings. This is consistent with the expectation that concentrations should be 

higher nearer to where BDEs are released from their use in commercial products―such as near a 

source burning such products (like the UK bonfire described above) or shredding/recycling such 

products (like the autoshredding and recycling sites studied by CARB and described above). 

(4) Comparisons with measurements in Europe and Asia are hard to make. One study from 

China suggests airborne concentrations well into the thousands, while other measurements in 

Europe and Asia appear more in line with those in the United States, with total BDEs mostly 

under 100 pg/m3

. 

4.5.2. Indoor Air Concentrations and Simultaneous Indoor/Outdoor Monitoring

 The primary sources of PBDEs to which the general population is exposed are the 

products in which they are used. This would suggest that indoor exposures would be of primary 

importance for this class of persistent and bioaccumulative toxics (PBTs), and, specifically, 

indoor air might contain higher concentrations of PBDEs as compared to outdoor air. Because 

the primary sources tend to be indoors, several researchers have focused their efforts in the 

indoor environment, and their efforts include simultaneous indoor/outdoor air measurements. In 

all these studies, the indoor measurements are higher than nearby outdoor measurements, often 
4-51

by factors of 10 or more. There is some evidence that combustion of products containing PBDEs 

release the PBDEs, but the evidence for this was primarily an open-air bonfire, not a controlled 

waste combustion process (see discussion above on the United Kingdom bonfire study by Farrar 

et al., 2004). There is also evidence that combustion of products containing PBDEs can result in 

formation of polybrominated dibenzo-p-dioxins and dibenzofurans (PBDD/Fs), and 

monobromo-polychlorinated dibenzo-p-dioxins and dibenzofurans (MoBPXDD/Fs). Hayakawa 

et al. (2004) found that the levels of MoBPXDD/Fs correlated positively with that of 

polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) in Japanese ambient air testing, 

and cited other laboratory testing showing concurrent formation of PCDD/Fs and MoBPXDD/Fs. 

They also showed a correlation between PBDD/Fs and some key PBDEs, citing other work 

showing formation for PBDFs from combustion of materials containing decaBDE. Thus, while 

the combustion of PBDEs appears to produce other brominated compounds, unlike dioxins, 

PBDEs are not formed themselves from combustion. 

 Unfortunately, U.S. researchers have not measured indoor air as much as house dust, or 

as much as researchers from Canada and abroad. Only three studies have been located that 

measured indoor air. The most recent involved sampling the air in 12 homes in Albany, NY, 

between December 2007 and January 2008 (Johnson-Restrepo and Kannan, 2009). They used a 

low-volume pump placed in the main room, the living room, and collected samples for 12 hours 

with an air flow of 25 L/min. Analysis of fractions on a glass fiber filter as well as a 

polyurethane foam plug allowed for collection of both the particle and vapor phases. The mean 

total concentration of 15 congeners (from BDE 28 to BDE 209) was 1.26 ng/m3, the median was 

0.76 ng/m3, and the range was 0.21 to 3.98 ng/m3. Unlike dust samples taken as part of the same 

study, BDE 209 could not be quantified at the detection limit of 0.1 ng/m3

, and the samples were 

dominated by BDE 47. 

The second was an indoor air study in which dust was sampled in 20 urban residences in 

the Boston, MA area, with the air data reported in Allen et al. (2007). At each home, one 

personal and two area samples were taken over the course of a week, presumably during 2005 

(the date of sampling was not provided, but the data were reported in 2006). Congeners include 

BDE 17, BDE 28/33, BDE 47, BDE 49, BDE 66, BDE 85/155, BDE 99, BDE 100, BDE 153, 

BDE 154, and BDE 209. The area samples were taken in the bedroom and living room, and the 

personal air sample was taken near the breathing zone in the bedroom. Participants turned on the 
4-52

samplers when returning in the evening and turned them off when leaving for work. The sums of 

the geometric mean concentrations of reported congeners over the 20 residences were 605, 392, 

and 366 pg/m3 for the personal, bedroom, and living room air. BDE 47 had the highest 

concentration, with geometric means of 227, 158, and 145 pg/m3 for the same three locations, 

with BDE 209 second highest at 174, 95, and 94 pg/m3, and BDE 99 third highest at 111, 67, and 

60 pg/m3

. 

 The third U.S. study found for indoor air measurement of PBDEs was conducted by 

CARB (2005). This study was also published in the open literature in Cahill et al. (2007). Air 

monitoring was conducted indoors within a computer training facility in a public office building, 

within and around an electronics recycling facility, and outside of an automotive shredding/metal 

recycling facility (the recycling and shredding sites were also described in the previous section) 

on the UC Davis campus. This was the most comprehensive single data set that could be found 

in the literature, with these four sites and the measurement of 33 congeners. The octa- and 

nonaBDEs were not measured, however, for most of the samples because there was analytical 

interference. The sampling for the indoor training facility included sampling when the 

computers were on versus off and, also, it included carpet dust samples. The dominant congener 

found in the air was BDE 47, followed by BDE 99, BDE 100, and BDE 28. In all cases, BDEs

were higher in air when the computers were on versus off, but the most meaningful difference 

was for BDE 209, when considered in terms of the relative difference between off and on, rather 

than on the magnitude of the concentration alone. Specifically, BDE 209 increased from 2 pg/m3

when the computers were off to concentrations from 56−74 pg/m3 when they were on in one set 

of tests and from 18, when off, to 47−56 pg/m3, when on, in the second test. The largest absolute 

concentration difference was for BDE 47, which was 213 and 112 pg/m3

 in the two tests (starting 

from concentrations in the 800-pg/m3 range in the off condition and then increasing by 213 and 

112 pg/m3 when the computers were turned on), followed by BDE 99, BDE 100, and then 

BDE 209. The average concentrations of total BDEs on (three measurements) averaged with off

(one measurement) on the 2 days of measurement were 1,550 and 2,010 pg/m3, with BDE 209 

averaging 50 and 65 pg/m3 for the 2 days. 

A second site studied by CARB was an electronics recycling facility. Outdoor 

measurements from two samplers in front and two in back suggested concentrations in the range 

of 2,000−3,000 pg/m3

, dominated by BDE 209. Generally, sampling indoors showed 
4-53

substantially higher BDE concentrations compared to outdoors at this facility. A total of six

samples were obtained from two samplers over 3 consecutive days in 2004. Concentrations 

ranged from 316,000 to 833,000 pg/m3. The overwhelmingly dominant congener was BDE 209, 

ranging from 79,000 to 833,000 pg/m3, with an average over 423,000 pg/m3. The next highest 

congeners were BDE 183, averaging 23,813 pg/m3, BDE 153 averaging over 5,600 pg/m3, 

BDE 154 averaging over 3,400 pg/m3

, and then BDE 99, BDE 47, and BDE 49, all at about 

1,770 pg/m3. Measurements of the filter and XAD-2 separately showed comparable amounts in 

vapor (XAD-2) as compared to particle (filter) for the tri BDEs, about 10 times more particle 

than vapor for the tetra BDEs, and then essentially none in the vapor phase for penta-, hexa-, 

hepta-, and marginal amounts for deca BDE congeners. 

 Only the sampling within the computer training facility showing total concentrations 

ranging from 1,500−2,000 pg/m3 might represent nonoccupational concentrations that could 

apply to the general public. However, these concentrations are high when compared to other 

indoor monitoring studies conducted in Canada and abroad. 

 Shoeib et al. (2004) reported on measurements from 10 indoor and 3 outdoor samplers in 

Toronto, Canada, using a traditional high-volume, two-phase air PS-1 air sampler. Samples were 

collected in November/December of 2001 and then again in March of 2003. Congeners included 

BDE 17, BDE 28/33, BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, and BDE 154. Total 

concentrations outdoors were 39 and 48 pg/m3 (two locations, three samples) and indoors were 

410, 358, 490, 2,088, 381, and 76 pg/m3

 (six locations, seven samples), leading to an average 

ratio difference of 15 between indoor and outdoor samples. BDE 47 represented about 46% of 

the sample, followed by BDE 99 at 25%. Both the Junge-Pankow model (Pankow, 1987) and 

the model based on the octanol air partition coefficient, the Koa model (Shoeib and Harner, 

2002), very well predicted the gas/particle partitioning that was observed in the high-volume

sampler. Observed particle phase percentages for the congeners, derived from the indoor 

sampler, from high to low are as follows: BDE 183―82%, BDE 154―82%, BDE 153―81%, 

BDE 85―75%, BDE 99―62%, BDE 66―24%, BDE 47―20%, BDEs 28/33―4%, and 

BDE 17―3%. 

Wilford et al. (2004) sampled air using a PUF disk as a passive sampler in 74 homes and 

at 7 outdoor sites during the winter of 2002/2003 in Ottowa, Canada. Indoor air concentrations 

of PBDEs were log-normally distributed, with a geometric mean of 120 pg/m3 (high of 
4-54

3,600 pg/m3), which is approximately 50 times higher than outdoor air concentrations, 

<0.1−4.4 pg/m3. Congeners measured include BDE 17, BDE 28, BDE 47, BDE 66, BDE 71, 

BDE 85, BDE 99, BDE 100, BDE 153, and BDE 154. The highest mean concentration was 

BDE 47 at 160 pg/m3, followed by BDE 99 at 42 pg/m3, and BDE 28 at 24 pg/m3. Wilford et al. 

(2004) stated that the indoor passive samplers are sampling mainly the vapor phase, with only a 

small contribution from particulates. The study found that the technical formulations tend to be 

enriched with heavier congeners, BDE 99 and BDE 100, while the indoor air was dominated by 

the lighter, more volatile congeners, BDE 47 and BDE 28. 

Limited studies abroad (United Kingdom and Kuwait) showed indoor concentrations of 

total BDEs similar to these Canadian studies, suggesting a range of 20−200 pg/m3. However, 

like the Canadian studies, measurements were not made for BDE 209. Hazrati and Harrad 

(2005) passively sampled 12 homes, 10 offices, and 1 private car in the United Kingdom for a 

period of 1 year, for sampling events that took between 4 and 6 weeks. Congeners sampled 

included BDE 28, BDE 47, BDE 49, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, and 

BDE 154. Concentrations ranged between 5 and 1,418 pg/m3

, with a mean of 148.2 and median 

of 38.4 pg/m3. In comparison with outdoor air from other studies, the authors suggested a 

20-fold difference between indoor and outdoor air. PUF samplers were used to measure air 

samples and bulk dust samples were obtained from vacuum bags in 17 homes in Kuwait between 

2/29/2004 and 4/11/2004 (Gevao et al., 2005). Individual congener concentrations included 

BDE 28, BDE 47, BDE 85, BDE 99, BDE 100, BDE 154, BDE 153, and BDE 183. Air 

concentrations ranged from 2.5−385 pg/m3, with a geometric mean of 10 pg/m3. BDE 47 was 

the most abundant congener representing, on average, 51% of the total PBDE concentration 

measured. The next most abundant congener, BDE 99, represented about 28% of the total. 

Harrad et al. (2004) presented congener data on BDE 47, BDE 99, BDE 100, BDE 153, 

and BDE 154 for indoor and outdoor air, and meat and vegan diets. The purpose of these 

measurements was to estimate daily exposure to BDEs via inhalation and diet. Indoor air 

concentrations were much higher than outdoor air concentrations; BDE 153 and BDE 154 were 

comparable in both environments, but BDE 47, BDE 99, and BDE 100 were each over 100 times 

higher indoors than outdoors. The mean total concentrations of BDEs in air in various 

environments include the following: 21 pg/m3

 in outdoor air, 525 pg/m3 in domestic indoor air, 

and 2,788 pg/m3

 in workplace environments. 
4-55

Harrad et al. (2006) collected indoor air samples at 92 locations in the United Kingdom, 

including 31 homes, 33 offices, 25 cars, and 3 public locations (post office, coffee shop, and 

supermarket) using PUFs. They measured BDE 28, BDE 47, BDE 49, BDE 66, BDE 85, 

BDE 99, BDE 100, BDE 153, and BDE 154. The overall average of all measurements was 

273 pg/m3

 total, with a median of 47 pg/m3 and a maximum of 8,180 pg/m3 found in cars. Cars 

had the highest average concentration, although that was skewed by three measurements greater 

than 2,000 pg/m3

; the highest concentration in the noncar environments was 1,416 pg/m3, and the 

average excluding the cars was 110 pg/m3. Like other studies, the major contributors to these 

concentrations were BDE 47 and BDE 99. 

Karlsson et al. (2007) reported on measurements of PBDEs in indoor air, house dust, and 

blood from five households, although air was not sampled in one of the households. Unlike 

some of these other studies, BDE 209 was measured. This congener was found in one of 

four samples, at 257 pg/m3 (DL = 173 pg/m3). BDE 47 was found in three of four samples, 

ranging from 126 to 171 pg/m3. The only other two congeners quantified in air were BDE 28 

and BDE 66, and they were found in seven of eight measurements ranging from 6 to 28 pg/m3

. 

Mandalakis et al. (2008) measured the levels of PBDEs inside automobile cabins in 

Greece. They measured the indoor air of 31 cars, taking 41 samples from low-volume samplers 

that were on for 48 hours, both when the car was in use and when not in use. The median 

concentration of total PBDEs was 201 pg/m3, dominated by BDE 209, explaining about half the 

total concentration. Seventeen other congeners were measured, with the next two predominant 

congeners being BDE 47 and BDE 99. The maximum found was over an order of magnitude 

higher than this median at 2,644 pg/m3. 

 In summary, only two studies could be located that measured indoor air concentrations of 

BDEs in the United States in residences rather than in occupational settings. One study in 

Albany, NY, showed a mean total concentration of 1,260 pg/m3

 and the other was of different 

rooms within urban residences in the Boston, MA, area, where the mean total household 

concentration was in the range of 200−500 pg/m3. The only other indoor measurements came 

from within a computer lab, and total concentrations were in the range of 1,500−2,000 pg/m3, 

with BDE 209 in the range of 50−70 pg/m3

. This was higher than indoor concentrations 

measured in Canada and abroad, which were mostly less than 200 pg/m3

 total, although 

BDE 209 was most often not measured in studies in Canada or abroad. Indoor 
4-56

industrial/occupational concentrations were substantially higher, with measurements in the 

hundreds of thousands of pg/m3. In United States, Canada, and studies abroad, indoor 

concentrations were found to be higher than outdoor concentrations in simultaneous 

measurements by factors of 10 to 100. BDE 47 appears to dominate indoor air concentrations, 

encompassing about one-third the total congener concentration when BDE 209 was measured in 

one study (Allen et al., 2007), and about half the total concentration in other studies when it was 

not measured. In the study where BDE 209 was measured in the indoor environment, it was the 

second highest concentration, followed by BDE 99. In a study of PBDE concentrations inside 

cars, BDE 209 dominated the profile, explaining about half the total concentration. 

4.6. FISH CONCENTRATIONS

This section reviews the data on fish, including fish caught in the wild, farmed fish, and 

fish samples from market basket surveys. While emphasis is on the U.S. studies, noteworthy 

studies from Canada and abroad are included as well. Like much of the data on other 

environmental levels, these studies often did not include BDE 183 and BDE 209, the primary 

markers for the octa and deca formulations, respectively. Another issue for evaluation and 

comparison of fish studies is that some results are reported on a wet weight (wwt) of whole 

tissue and others on a lipid-basis. Like other PBTs, PBDEs bioaccumulate in lipids of animals, 

so it would be useful to have all data reported on both wwt and lipid bases. While the authors 

reported results are provided below, wwt or lipid-based (lwt) concentrations are concurrently 

provided in parentheses when possible. Table 4-3 provides congener-specific fish concentrations 

for fish caught in the United States. 

4.6.1. Farmed Fish Concentrations

A total of 70 farmed and wild salmon were collected from wholesale and retail outlets in 

Maine in August 2003 and May 2004 (Shaw et al., 2008). They represented salmon from three 

regions: two farms in eastern Maine, three farms in eastern Canada, and one farm in Norway. 

Samples were composited so that the Maine sample results were displayed for the two farms. 
4-57

Table 4-3. Congener-specific concentrations of PBDEs for fish caught in the 

United States (units in ng/g wwt, or lwt, in parentheses if available)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

MonoBDE

1 0.007, 0.006

0.007, 0.006

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

2 0.005, 0.004

0.005, 0.004

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

3 0.004, 0.004

0.004, 0.004

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

DiBDE

7 0.001, 0.0002

0.001, 0.001

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

8/11 0.0002, 0.0002

0.0004, 0.0003

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

10 0.0002, 0.0002

0.0004, 0.0003

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

12/13 0.0003, 0.0002

0.0003, 0.0002

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

15 0.006, 0.003

0.002, 0.002

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

TriBDE

17 0.011 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring

Schecter et al. 

(2006a)

0.002 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%

Schecter et al. 

(2009b)

0.10 (0.77 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites (2004)

0.010, 0.003

0.003, 0.003

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

25 0.007, 0.002

0.001, 0.001

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

28 0.01 Average of 18 farm and wild-caught salmon from 

Maine, Alaska, Canada; eyeball estimate from graph 

Shaw et al. (2008)

0.01 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%

Schecter et al. 

(2009b)

0.026 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring

Schecter et al. 

(2006a)

1.67 (10.21 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites (2004)
4-58

Table 4-3. Congener specific concentrations of PBDEs for fish caught in the 

United States (units in ng/g wwt, or lwt, in parentheses if available)

(continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

1.2 (5.4 lwt) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

28/33 0.036, 0.014

0.006, 0.006

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

30 0.0008, 0.0007

0.0008, 0.0006

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

32 0.0007, 0.0006

0.0007, 0.0005

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

35 0.005, 0.003

0.003, 0.0006

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

37 0.001, 0.001

0.0005, 0.0004

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

TetraBDE

47 0.60 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring

Schecter et al. 

(2006a)

0.20 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%

Schecter et al. 

(2009b)

3.6 n = 17 locations in SF Estuary; bivalve (clam, oyster, 

mussel) average

Oros et al. (2005)

20.6, 10.9 Means from 2004 (n=132) and 2005 (n=213) from 

national study on rivers and streams in the US

Blocksom et al. 

(2010)

0.4 Average of 18 farm and wild-caught salmon from 

Maine, Alaska, Canada; eyeball estimate from graph

Shaw et al. (2008)

52.1 Average of 21 salmon samples from Lake Michigan Manchester-Neesvig 

et al. (2001)

63.72 (391.12 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)

2−110 Time/spatial trend study for lake trout/smelt in Great 

Lakes; range for last samples, ~2003

Batterman et al.

(2007)

46.0 (488.3 lwt) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

1.5, 22, 84% n = 63; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

2.6, 1.9

0.1, 0.1

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

49 4.78 (35.35 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)
4-59

Table 4-3. Congener specific concentrations of PBDEs for fish caught in the 

United States (units in ng/g wwt, or lwt, in parentheses if available)

(continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.059 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%

Schecter et al. 

(2009b)

ND, 1.3, 33% n = 60; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

0.1, 0.03

0.01, 0.01

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

66 1.82 (11.29 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)

0.005 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 2 

detects

Schecter et al. 

(2009b)

0.021 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring

Schecter et al. 

(2006a)

ND Average of 18 farm and wild-caught salmon from 

Maine, Alaska, Canada; eyeball estimate from graph

Shaw et al. (2008)

1.7 Average of 21 salmon samples from Lake Michigan Manchester-Neesvig 

et al. (2001)

0.3 (1.8 lwt) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

ND, 1.0, 19% n = 27; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

0.03, 0.02

0.002, 0.001

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

71 ND Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)

ND, <0.5, 3% n = 63; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

75 0.008, 0.008

0.001, 0.001

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

77 0.001 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring; mostly NA or ND, with one 

3.6 herring

Schecter et al. 

(2006a)

0.001, 0.0005

0.0008, 0.0005

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

PentaBDE

85 0.67 (4.31 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)
4-60

Table 4-3. Congener specific concentrations of PBDEs for fish caught in the 

United States (units in ng/g wwt, or lwt, in parentheses if available)

(continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.004 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring; mostly NA or ND; one catfish at 

41.6

Schecter et al. 

(2006a)

0.0006 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 

1 detect at 0.002

Schecter et al. 

(2009b)

ND Average of 18 farm and wild-caught salmon from 

Maine, Alaska, Canada; eyeball estimate from graph

Shaw et al. (2008)

0.1 (0.9 lwt) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

0.03, 0.01

0.003, 0.003

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

88 0.03 Average of 2 farm salmon farm composited from 

Maine; eyeball estimate from graph 

Shaw et al. (2008)

99 22.96 (142.20 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)

0.4−15 Time/spatial trend study for trout/smelt in Great 

Lakes; range for last sampled, ~2003

Batterman et al. 

(2007)

0.17 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring

Schecter et al. 

(2006a)

0.04 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%

Schecter et al. 

(2009b)

0.1 Average of 18 farm and wild-caught salmon from 

Maine, Alaska, Canada; eyeball estimate from graph

Shaw et al. (2008)

9.3 Average of 21 salmon samples from Lake Michigan Manchester-Neesvig 

et al. (2001)

1.2 n = 17 locations in SF Estuary; bivalve (clam, oyster, 

mussel) average

Oros et al. (2005)

7.9, 2.7 Means from 2004 (n=132) and 2005 (n=213) from 

national study on rivers and streams in the US

Blocksom et al. 

(2010)

5.5 (23.1 lwt) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

ND, 17.0, 38% n = 63; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

1.6, 0.9

0.1, 0.1

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

100 17.50 (109.51 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)
4-61

Table 4-3. Congener specific concentrations of PBDEs for fish caught in the 

United States (units in ng/g wwt, or lwt, in parentheses if available)

(continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.5−15 Time/spatial trend study for trout/smelt in Great 

Lakes; range for last sampled, ~2003

Batterman et al. 

(2007)

0.13 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring

Schecter et al. 

(2006a)

0.04 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%

Schecter et al. 

(2009b)

0.08 Average of 18 farm and wild-caught salmon from 

Maine, Alaska, Canada; eyeball estimate from graph

Shaw et al. (2008)

9.7 Average of 21 salmon samples from Lake Michigan Manchester-Neesvig 

et al. (2001)

0.9 n = 17 locations in SF Estuary; bivalve (clam, oyster, 

mussel) average

Oros et al. (2005)

4.8, 2.0 Means from 2004 (n=132) and 2005 (n=213) from 

national study on rivers and streams in the US

Blocksom et al. 

(2010)

14.0 (55.4) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

1.0, 5.1, 51% n = 63; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

0.6, 0.4

0.03, 0.03

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

116 0.004,0.003

0.002, 0.001

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

119 0.009, 0.007

0.0008, 0.0006

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

0.002 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 

2 detects

Schecter et al. 

(2009b)

126 0.002, 0.001

0.00080.0006

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

HexaBDE

138 ND Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)

0.001 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring; mostly ND except 2 catfish at 

5.1 and 7.9

Schecter et al. 

(2006a)

0.0007 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 

2 detects

Schecter et al. 

(2009b)
4-62

Table 4-3. Congener specific concentrations of PBDEs for fish caught in the 

United States (units in ng/g wwt, or lwt, in parentheses if available)

(continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

ND Average of 18 farm and wild-caught salmon from 

Maine, Alaska, Canada; eyeball estimate from graph

Shaw et al. (2008)

<0.001 Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

ND, <0.9, 2% n = 63; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

0.01, 0.003

0.001, 0.001

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

153 0.021 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring; 

Schecter et al. 

(2006a)

0.006 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%

Schecter et al.

(2009b)

4.43 (27.48 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)

1.2, 0.3 Means from 2004 (n=132) and 2005 (n=213) from 

national study on rivers and streams in the US

Blocksom et al. 

(2010)

0.1−3.6 Time/spatial trend study for trout/smelt in Great 

Lakes; range for last sampled, ~2003

Batterman et al. 

(2007)

ND Average of 18 farm and wild-caught salmon from 

Maine, Alaska, Canada; eyeball estimate from graph

Shaw et al. (2008)

3.8 (17.0 lwt) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

ND, 1.1, 40% n = 63; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

0.3, 0.2

0.01, 0.01

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

154 0.049 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring

Schecter et al. 

(2006a)

0.015 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%

Schecter et al. 

(2009b)

7.63 (46.84 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)

0.9, 0.3 Means from 2004 (n=132) and 2005 (n=213) from 

national study on rivers and streams in the US

Blocksom et al. 

(2010)

ND Average of 18 farm and wild-caught salmon from 

Maine, Alaska, Canada; eyeball estimate from graph

Shaw et al. (2008)

9.2 (2.6 lwt) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)
4-63

Table 4-3. Congener specific concentrations of PBDEs for fish caught in the 

United States (units in ng/g wwt, or lwt, in parentheses if available)

(continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.48, 0.88, 49% n = 63; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

0.2, 0.2

0.01, 0.01

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

155 0.2, 0.2

0.01, 0.01

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

156 0.003, 0.003

0.002, 0.002

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

166 0.003, 0.003

0.002, 0.002

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

HeptaBDE

181 0.002, 0.002

0.002, 0.002

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

183 0.002 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring

Schecter et al. 

(2006)

0.001 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 

1 detect at 0.004

Schecter et al. 

(2009b)

0.13 (0.73 lwt) Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)

0.02, 0.01 Means from 2004 (n=132) and 2005 (n=213) from 

national study on rivers and streams in the US

Blocksom et al. 

(2010)

1.1 (3.1 l2t) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

ND, <0.9, 3% n = 63; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

0.02, 0.01

0.003, 0.003

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

184 ND, <0.9, 2% n = 60; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

190 ND Time/spatial trend study for lake trout in several Great 

Lakes; average for 2000 only

Zhu and Hites 

(2004)

0.003, 0.003

0.003, 0.002

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

191 ND, <0.9, 0% n = 60; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)
4-64

Table 4-3. Congener specific concentrations of PBDEs for fish caught in the 

United States (units in ng/g wwt, or lwt, in parentheses if available)

(continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

196 ND n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 

DL < 0.003

Schecter et al. 

(2009b)

197 0.009, 0.008

0.002, 0.002

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

0.0002 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 

1 detect at 0.001

Schecter et al. 

(2009b)

OctaBDE

203 1.3 (4.3 lwt) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

0.003, 0.002

0.003, 0.002

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

NonaBDE

206 0.0004 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 

1 detect at 0.003

Schecter et al. 

(2009b)

207 0.007, 0.005

0.005, 0.005

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)

0.0007 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 

1 detect at 0.005

Schecter et al. 

(2009b)

DecaBDE

209 0.092 n = 24; consumer fish including tuna, salmon, shark, 

trout, catfish, herring; majority ND but one catfish at 

1,269

Schecter et al. 

(2006a)

0.008 n = 8, consumer fish including salmon, tuna, catfish, 

canned sardines, frozen fish sticks, lipid = 8.7%; 

2 detects at 0.007 and 0.051.

Schecter et al. 

(2009b)

0.5 (96.6 lwt) Average of 88 specimens, 9 species of marine fish, 

2 species of dolphin, off Florida coast

Johnson-Restrepo et 

al. (2005)

ND, <5.3, 6% n = 63; median, mean, frequency detection in 

Washington State survey in rivers and lakes

Johnson et al. (2006)

0.1, 0.04

0.1, 0.08

Mean, median of wild-caught (n = 33) and farmed 

salmon (n = 28) in southern Mississippi

Staskal et al. (2008)
4-65

 Results were provided for nine congeners including BDE 28, BDE 47, BDE 66, BDE 85, 

BDE 99, BDE 100, BDE 138, BDE 153, and BDE 154. The samples were analyzed with skin on 

and skin off to see if that made a difference; the results suggested essentially no difference. 

There was also no correlation between lipid content and wwt concentrations, which is 

counterintuitive. It was found that total concentrations between the farmed and wild-caught were 

essentially identical, and the values ranged between 0.42 and 1.37 ng/g total wwt. This is four to 

five times lower than PBDE concentrations reported in farmed salmon from British Columbia 

and northwestern Europe, but comparable to levels found in farmed salmon from Chile. BDE 47 

dominated the profiles, contributing 38−58% of the total measured PBDEs, with BDE 99 plus 

BDE 100 accounting for another 21−26% of the total. 

Jacobs et al. (2002) measured PCBs, dichlorodiphenyltrichloroethane (DDT), and PBDEs 

in farmed and wild European Atlantic salmon, aquaculture feeds, and fish oils used to 

supplement the feed. Seven British salmon samples, five additional salmon samples (two from 

Ireland, and three purchased from a Belgian market), eight salmon feeds (from four different 

Scottish sources), five fish oils, and one vegetable oil were analyzed. Congeners analyzed 

included BDE 28, BDE 71, BDE 47, BDE 75, BDE 66, BDE 100, BDE 99, BDE 153, and 

BDE 154. Total BDEs ranged from 1.1−85.2 ng/g lwt in 13 salmon samples (average = 33.8), 

with the highest found in a wild salmon sample. BDE 47 predominated, averaging about 53% of 

total. The levels of BDE in feed ranged from 8.1 to 23.9 ng/g lwt for eight feed samples, and the 

range in fish oil was ND to 12.7 ng/g lwt. BDE 47 similarly dominated the feed and fish oil 

samples.

Staskal et al. (2008) collected 60 samples of catfish from southern Mississippi, including 

28 that were farm-raised and store-bought, and 33 that were wild-caught. Samples were 

analyzed for 43 congeners including BDE 209. The wild-caught catfish were significantly 

higher (p < 0.01) than the store-bought fish: the median total concentration (sum of 

43 congeners) from the wild-caught fish was 10.6 ng/g wwt (1,670.4 ng/g lwt), while the median 

for the farm-raised catfish was 0.5 ng/g wwt (5.8 ng/g lwt). BDE 47, BDE 99, BDE 100, 

BDE 153, and BDE 154 were detected in all samples and dominated the concentrations. 

BDE 209 was detected in 30% of the samples, and the average concentration in samples above 

the LOD was 0.2 ng/g wwt. BDE 209 was detected in the three catfish nugget samples of the 
4-66

study and contributed about 36% to the total concentration in these nuggets. Congener-specific 

data from this study are provided in Table 4-3. 

 The most comprehensive study on BDEs in farmed fish was conducted by Hites et al. 

(2004). PBDEs were measured in about 700 farmed and wild salmon collected from around the 

world, including from Maine and Washington, at the AXYS Analytical Lab in Sidney, BC. 

BDE 1, BDE 2, BDE 3, BDE 7, BDE 8, BDE 10, BDE 11, BDE 12, BDE 13, BDE 15, BDE 17, 

BDE 25, BDE 28, BDE 30, BDE 32, BDE 33, BDE 35, BDE 37, BDE 47, BDE 49, BDE 66, 

BDE 71, BDE 75, BDE 77, BDE 85, BDE 99, BDE 100, BDE 105, BDE 116, BDE 119, 

BDE 126, BDE 138, BDE 140, BDE 153, BDE 154, BDE 155, BDE 166, BDE 191, BDE 183, 

BDE 190, BDE 206, BDE 207, and BDE 208 were found. BDE 209 was tested for but not 

detected with a detection limit of 0.1 ng/g wwt (other congeners had DLs at 0.001 to 

0.01 ng/g wwt). The concentrations of total PBDEs in these samples ranged up to 10 ng/g wwt. 

In order of magnitude, the samples suggested the following: Europe farmed salmon > North 

America farmed salmon > Chile farmed salmon > wild salmon. Farmed salmon ranged from 

around 0.3 up to nearly 10 ng/g wwt, while the bulk of wild salmon were less than 0.3 ng/g wwt. 

In terms of wild salmon species, Chinook was the highest, with an average of 2.26 ng/g wwt for 

the 9 samples as compared to 0.13 ng/g wwt for the 36 others. In terms of composition, BDE 47 

dominates (around 47%), followed by BDE 99 (18%), BDE 100 (10%), BDE 28/33 and BDE 49 

(5%), with other congeners under 5%. Although not found in the farmed salmon, BDE 209 was 

found in the salmon feed and at about 15% of total concentration. Total PBDE concentrations in 

13 feed samples ranged from 0.5 to 10.9 ng/g wwt. 

Hayward et al. (2006) collected 18 samples of farmed and wild-caught fish from 

supermarkets in Maryland or Washington, DC in 2004, and an additional four samples were 

collected in 2001, including one from North Carolina. There were 6 farm-raised salmon, 

6 wild-caught salmon, 5 bluefish, and 5 rockfish among the 22 samples. They were analyzed for 

22 BDEs, including BDE 209, but only BDE 28, BDE 47, BDE 49, BDE 99, BDE 100, 

BDE 153, and BDE 154 could be found routinely. Individual congener results were graphed and 

not provided in a table, so they were not added to Table 4-3. However, they appear consistent 

with other measurements described here and in Table 4-3. Wild bluefish had the highest total 

PBDE concentration, averaging 15.1 ng/g wwt (n = 5), followed by wild rockfish at 

5.4 ng/g wwt (n = 5), farmed salmon at 1.0 ng/g wwt (n = 6), and Alaskan salmon (king, coho, 
4-67

and sockeye) at 0.4 ng/g wwt (n = 6). BDE 47 had the highest concentrations, explaining 

55−70% of the total, followed by BDE 99 and BDE 100, both of which took turns being second 

behind BDE 47 in different samples. BDE 153 was not found in wild salmon and only found at 

1−3% of total in other species, and BDE 183 was found in only one sample. 

4.6.2. Freshwater and Marine Fish Concentrations

Fish from the major river systems of the central US - the Upper Mississippi, Missouri, 

and Ohio Rivers - were sampled as part of EPA’s Environmental Monitoring and Assessment 

Program for Great River Ecosystems (EMAP-GRE; Blocksom et al., 2010). The overall design 

was a probability design aimed at characterizing these systems on a geographic basis. A first set 

of results from this study pertaining to sampling years 2004 and 2005 included overall results for 

legacy organochlorines including pesticides and PCBs as well as emerging compounds of 

concern, including PBDEs. Both “large” and “small” fish were sampled and composited per 

“site” of sampling. The large species represent fish targeted by recreational anglers and could 

include bottom (catfish, e.g.) as well as column fish (largemouth bass, e.g.). The small fish were 

meant to be representative of prey for piscivorous wildlife. For purposes here, the large fish 

PBDE results were of interest. A total of 345 large samples (132 in 2004 and 213 in 2005) were 

measured for 6 PBDEs, including BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 

183. Individual congener concentrations were not presented in Blocksom et al (2010), but were 

provided by personal communication (JM Lazorchak, US EPA, 26 W. Martin Luther King Dr 

Cincinnati, OH 45268). The mean concentrations of all congeners (calculated assuming ND = 0, 

DL = 1 ng/g wwt) each year are provided in Table 4.1. Overall, BDE 47 was detected the most 

frequently (nearly 100% of the time) at the highest average concentration, 20.6 ng/g wwt in 2004 

and 10.9 ng/g wwt in 2005. In contrast, BDE 183 was detected less than 1% of the time, and 

when detected, at concentrations near 1 ng/g wwt. The total concentration (sum of the 6 

congeners) averaged 35.4 ng/g wwt in 2004 and 16.2 ng/g wwt in 2005. 

Oros et al. (2005) measured 22 BDE congeners but detected only BDE 47, BDE 99, and 

BDE 100 in bivalves (clams, mussels, and oysters) in composited samples originating from 

16 locations, at wwt concentrations with an average of 5.7 ng/g wwt, ranging from 

2−13 ng/g wwt (855−13,502 ng/g lwt) in the San Francisco Estuary. The congener found most 
4-68

abundantly was BDE 47; it is found in all samples above its DL and comprising about 50−70% 

of the total concentration, while BDE 99 and BDE 100 were found in about 50% of the samples.

Manchester-Neesvig et al. (2001) sampled 21 coho and Chinook salmon from Lake 

Michigan tributaries in 1996 and analyzed them for six PBDEs: BDE 47, BDE 66, BDE 99, 

BDE 100, BDE 153, and BDE 154. These samples are among the highest fish (and terrestrial 

animal) concentrations in the literature, ranging from 45 to 148 ng/g wwt (773 to 

8,120 ng/g lwt), with an average of 80.1 ng/g wwt (2,440 ng/g lwt). BDE 47 dominated the 

concentration, comprising 56% of the total, followed by BDE 99 at 19%, BDE 100 at 12%, 

BDE 154 at 6.6%, and BDE 153 at 3.6%. 

A more comprehensive evaluation of Great Lakes fish was conducted by Zhu and Hites 

(2004). Lake trout collected between 1980 and 2000 from Lakes Superior, Michigan, Huron, 

and Ontario, and walleye from Lake Erie, were analyzed for 15 congeners, including BDE 17, 

BDE 28, BDE 47, BDE 49, BDE 66, BDE 71, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, 

BDE 154, BDE 183, BDE 190, and BDE 209. The levels of BDE 209 found in all fish samples 

were in the undetectable range of 3.6 ng/g lwt, thus it was deemed to be not detected in all 

samples and not reported. Total PBDEs (sum of BDE 47, BDE 99, BDE 100, BDE 153, and 

BDE 154 only) rose from under 15 ng/g lwt in 1980 to the range of 58−180 ng/g lwt in 1990, and 

finally in the range of 400−1,400 ng/g lwt in 2000. The authors concluded that the trends 

suggest a doubling time of 3−4 years. The authors discussed some spatial trends, but mostly the 

concentrations in the lakes over time were comparable. The summary information is the average 

congener concentration over all lakes for the most recent year, 2000. There was a systematic 

change of congener distribution over time. More of the octaBDE product, which was high in 

BDE 153 and BDE 154, was used in the 1980s, and there was a higher proportion of these two in 

the 1984 fish as compared to the 1996 fish. The ratio of pentaBDE use divided by octaBDE in 

the mid-1980s was 0.7 (more octa used). However, by the mid-1990s, there was a shift, and the 

ratio was now 2 (twice as much penta used). Subsequently, fish in the latter 1990s were

dominated by BDE 47. 

Batterman et al. (2007) also conducted a comprehensive temporal study of archived and 

fresh fish from the Great Lakes. Chapter 3 summarized this study as it measured numerous 

archived and freshly caught fish from all of the Great Lakes from 1979 to the present. Batterman 

et al. (2007) focused only on BDE 47, BDE 99, BDE 100, and BDE 153 and found consistent 
4-69

increases, with doubling times in the range of 2−4 years. The total concentrations (sum of the 

four congeners) in trout ranged from about 20 to over 100 ng/g wwt in the last samples taken, 

between 2000 and 2005. Concentrations were a bit lower in the last samples of smelt taken, 

between 3 and 81 ng/g wwt. BDE 47 dominated the total, contributing about 70% of total 

concentration, with BDE 99 and BDE 100 contributing about 13% each. 

 Although the focus was on analog compounds (which are defined as compounds 

structurally analogous to PBDEs, such as hydroxylated PBDEs, or OH PBDEs), six congeners, 

including BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183 were measured in the 

plasma of fish from the Detroit River (Valters et al., 2005). Composited plasma samples (blood 

was centrifuged to separate red blood cells from plasma) were obtained from 13 fish species. 

Total PBDEs ranged from 0.16 to >21.1 ng/g wwt (in channel catfish; all other samples less than 

10.0 ng/g wwt). Given that plasma was 0.6−2.7% lipids, lipid concentrations would be much 

higher, at over 200 ng/g lwt. BDE 47 was the dominant congener (over 60% in most samples), 

followed by BDE 99 and BDE 100, which were of similar magnitude; then BDE 153 and 

BDE 154 at lower and sometimes nondetect values. BDE 183 was only detected in channel 

catfish at 0.805 ng/g wwt; BDE 47 was detected at 11.5 ng/g wwt in this catfish. 

A total of 63 samples of fish were collected from Washington State rivers and lakes by 

the Washington State Department of Ecology in 2005 and 2006 (Johnson et al., 2006). These 

samples were measured for 13 congeners including BDE 47, BDE 49, BDE 66, BDE 71, 

BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, BDE 184, BDE 191, and 

BDE 209. Results for total ranged from not detected to 1,059 ng/g wwt, with a median of 

2.8 ng/g wwt and a mean of 35 ng/g wwt (derived assuming ND = 0). The most frequently 

detected congener was BDE 47; it was detected 84% of the time, with a median concentration of 

1.5 ng/g wwt. It accounted for 68% of the total concentration found, with BDE 99 accounting 

for 16% as the second highest concentration. Other congeners detected frequently equal 

BDE 100, BDE 154, BDE 153, and BDE 99, detected 51%, 49%, 40%, and 38% of the time, 

respectively. BDE 209 was not detected, but it had a high detection limit, ranging from 

1−6 ng/g wwt (all other detection limits were less than 0.5 ng/g wwt). The Spokane River was 

the most impacted river, with three samples taken measuring 76, 417, and the survey maximum 

of 1,059 ng/g wwt. This river also had the highest water concentrations in this same survey, with 

two measurements above 100 pg/L at 146 and 926 pg/L, with all other water measurements well 
4-70

under 100 pg/L. Johnson et al. (2006) reported on an earlier study done by the Washington State 

Department of Ecology on fish in the Spokane River, and they similarly found high 

concentrations, ranging from 30 to 1,222 ng/g wwt in 47 samples (some of which were 

composites). 

 Fish from three locations in the Savannah River were sampled in 2005 and analyzed for 

12 BDEs (Sajwan et al., 2006). Individual congener results were not provided, and results 

suggested generally higher levels of BDEs, similar to other freshwater bodies, with totals at 10 to

>300 ng/g lwt (which would equal about 1/10 as much on a fwt basis). The interesting finding 

from this study was that the predominant congener was identified as BDE 30 in all three 

locations―a congener that was not looked for in any other fish study. The second most 

predominant congeners were the familiar BDE 47 and BDE 99. 

 Ashley et al. (2006) analyzed samples of eel and sediment from the Delaware River 

(sediment samples described in Section 4.2). The eel samples were collected in 1998 as part of 

an earlier effort focusing on PCBs. Eel was selected as a good bioindicator of the quality of the 

water body since eels have a small range of habitat throughout the water body where they live. 

Like other aquatic biotal samples from freshwater systems, the PBDE concentrations in these 

eels were high, with totals ranging from 1 to 408 ng/g wwt, with an average of 86 ng/g wwt over 

17 samples. The predominant congener was BDE 47, explaining 56% of the concentrations. 

BDE 100 was the next most predominant congener at 30%, and other congeners contributed 3% 

or less. BDE 209 was not detected in the samples, although the detection limit was not provided. 

Hale et al. (2001) measured BDE 47, BDE 49, BDE 99, BDE 100, BDE 153, and 

BDE 154 in 70 samples of fish from the Roanoke River, the Dan River, and the Hyco River, 

which are within two large watersheds in Virginia. Overall statistics of findings were not 

provided. It was noted that BDE 47 was quantified (at over 5 ng/g lwt) in 88% of samples, with 

over half the fish samples having concentrations greater than 100 ng/g lwt, and with 16 samples 

containing BDE 47 at over 1,000 ng/g lwt. The highest total concentration was 47,900 ng/g lwt. 

Overall, BDE 47 dominated the profiles, explaining between 40 and 75% for the five different 

species tested, with BDE 99 and BDE 100 explaining in the range of 5−20% of concentrations. 

BDE 49 was also measured and identified, generally explaining less than 5% of total PBDE 

concentrations in the sampled fish. 
4-71

Marine fish sampled off the coast of Florida were analyzed for 12 PBDEs by 

Johnson-Restrepo et al. (2005). Specifically, a total of 88 specimens from nine species of marine 

fishes and two species of dolphins were analyzed for BDE 28, BDE 47, BDE 66, BDE 85, 

BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, BDE 203, and BDE 209. The 

overall mean concentrations of total PBDEs were 10-fold greater than the levels reported for 

Arctic marine fish such as polar cod from Storfjorden, and for several fish species from southern 

Greenland. On the other hand, the concentrations were one order of magnitude lower than what 

was reported for lake trout in the Great Lakes. Total concentrations in teleost fishes ranged from 

8 ng/g lwt to 88 ng/g lwt, with a mean of 43 ng/g lwt. Mean concentrations of total BDEs in 

sharks and dolphins were higher, ranging from 37.8 ng/g lwt to 1,630 ng/g lwt. BDE 47 was 

overwhelmingly found at the highest concentrations in teleost fish and dolphins (10s of ng/g lwt 

in teleost, but 100s of ng/g lwt in dolphins), but BDE 209 was the highest found in sharks 

(16−778 ng/g lwt). Other than for sharks, BDE 209 was found at an average of 0.5 ng/g lwt, 

with several nondetects. Johnson-Restrepo et al. (2005) also did a temporal study, evaluating 

nine samples of dolphin blubber collected between 1991 and 1996, with six analogous samples 

collected between 2001 and 2004. The authors similarly compared bull shark samples collected 

in 1993 and 1994 with those collected in 2002 and 2004. A clear elevation was seen in both 

PCB and PBDE concentrations. Mean total PBDEs in the dolphins rose from 363 to 

1,190 ng/g lwt, and mean total PBDEs in sharks rose from 78 to 1,630 ng/g lwt. 

 Rayne et al. (2003) studied the temporal trend of increasing BDE concentrations in fish 

from the Columbia River System in southeastern British Columbia. The concentrations of 

33 individual mono- through hexabrominated BDE congeners (which includes congeners up to 

BDE 156) were measured in mountain whitefish, large scale suckers, and surficial sediments 

from several locations on the Columbia and Kootenay River systems in southeastern British 

Columbia, Canada. A total of 41 whitefish samples were obtained from the period of 1992 to 

2000, specifically 1992, 1994 and 1995, 1998, and 2000. Eleven sediment and six sucker fish 

samples were taken in 2001 and 2000, respectively. Total PBDE concentrations in whitefish , 

increased by a factor of 11.8 at one location over the period from 1992 and 2000: they went from 

6.1 ng/g wwt (average) in 1992 to 19.1 ng/g wwt in 1994 and 1995 to 71.8 ng/g wwt in 2000. At 

a second location, total PBDE concentrations in whitefish increased by a factor of 6.5: the 

starting concentration of 4.5 ng/g wwt in 1992 rose to 29.2 ng/g wwt by 2000. At a remote site, 
4-72

total PBDE was 0.9 ng/g wwt in the whitefish. The authors attributed the higher concentrations 

seen at the other two sites to domestic wastewater draining directly into the river. Sucker fish

sampled in 2000 had lower concentrations than whitefish at 5.0 ng/g wwt. The sediment 

congener pattern was somewhat different from the fish. The primary congeners in the sediment 

and fish were BDE 47, BDE 99, and BDE 100, but BDE 47 was the major congener in sediments 

(46−63% total), followed by BDE 99 (23−39%), and BDE 100 (6−8%), while for whitefish, 

BDE 99 was the dominant congener, followed by BDE 47 and BDE 100. This is contrary to 

other literature, which showed similar dominance of BDE 47 in both sediment and fish. 

Vives et al. (2004) collected fish from 11 mountain lakes in Europe and 1 in Greenland. 

The importance of these lakes is that the only way BDEs could have reached these aquatic 

ecosystems is by long-range transport. Liver and muscle tissue of trout (brown trout, brook 

trout, and Arctic char) were sampled. The authors were unclear as to which BDEs were 

measured. They stated that all congeners were identified using external standards composed of 

39 individual congeners up to BDE 190 (not including BDE 209), but they only presented results 

for BDE 28, BDE 33, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154, while not providing 

any information on other congeners, including those that they looked for but did not detect. 

Results from 55 trout specimens were as follows: 0.1−1.3 and 0.07−0.77 ng/g wwt in liver and 

muscle, respectively (2.4−40.0 and 2.9−41.0 ng/g lwt). BDE 47 was the highest congener found, 

followed by BDE 99, BDE 100, BDE 153, BDE 154, and BDE 28. This pattern was much more 

apparent in liver as compared to muscle. There was an age relationship found in the data: older 

fish had significantly higher BDE concentrations. Concentration increases of between 4 and 12 

times were found between 1- and 21-year-old individual fish samples. 

 Peng et al. (2005) collected 60 tissue samples from six rivers and three estuaries in 2003 

in China. BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183 were 

measured. Results showed river-average results ranged from 25 to 152 ng/g lwt, and 31 to 

281 ng/g lwt from the estuaries. The fish species were not identified. In all rivers, BDE 47 was 

the predominant congener found, exceeding other congeners by factors of 3 or more and 

comprising over 50% of the total concentration. In most cases, BDE 154 was the next highest, 

followed by BDE 99 and BDE 100. The one exception was an estuary, where BDE 154 

accounted for 101 ng/g lwt of a total of 281 ng/g lwt, followed by BDE 47 at 92 ng/g lwt. In all 

other locations, BDE 154 was 10 or less ng/g lwt. 
4-73

4.6.3. Fish from the Retail Marketplace

 The fish data that might be considered the most relevant for estimating exposure to the 

general population are from the retail market basket sampling described in this section. Also, 

this was essentially the only data on fish in the literature that reported measurements of BDE 183 

and BDE 209. Fish were either the primary target of the sampling or included among a variety 

of food products. 

Schecter et al. (2006a; derived from data in Schecter et al., 2004, with additional data and 

analyses) reported the results of a comprehensive market basket survey entailing 62 samples that 

encompassed meat products, dairy products, and fish. BDE 17, BDE 28, BDE 47, BDE 66, 

BDE 77, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, and BDE 209 

were reported, although the average results reported, which were based on the substitution of 1/2 

DL for ND, could be misleading because there was a high frequency of ND (not detected) or NA

(not analyzed) for many congeners. For example, BDE 17 was only detected in 4 of 18 samples, 

at detection limits ranging from 0.0001 to 0.0007 ng/g wwt. However, the mean was reported as 

0.0008 ng/g wwt, due to several samples with detection limits above 0.002 ng/g wwt. The 

detection limit issue was even more apparent with BDE 209. For example, the mean 

concentration for BDE 209 in fish was 0.1 ng/g wwt, but it was not detected in 14 of 24 samples. 

The high mean was driven by two samples at 1.27 ng/g wwt and 0.68 ng/g wwt; all other positive 

detections were less than 0.025 ng/g wwt. The detection limits on BDE 209 were highly 

variable, ranging from 0.011 ng/g wwt to 0.17 ng/g wwt. The data regarding the more 

commonly found BDEs at higher concentrations (congeners BDE 47, BDE 99, BDE 100, 

BDE 153, and BDE 154) must be assumed to be valid. A second key limiter of this data set is 

that all samples were purchased from supermarkets in Dallas, TX, so it is unclear to what extent 

these data represent national trends. A large proportion of the food supply is national in scope, 

such as nonperishable or frozen food items, but others might be locally produced (fruits and 

vegetables) or perhaps regionally produced (animal food products). Still, this data set represents 

one of the most comprehensive surveys of terrestrial animal food products in the United States

analyzed for PBDEs. The top four BDE congeners found in fish were BDE 47 at 0.6 ng/g wwt, 

BDE 99 at 0.17 ng/g wwt, BDE 100 at 0.13 ng/g wwt, and BDE 209 at 0.098 ng/g wwt. The 
4-74

mean concentration of BDE 183, the primary marker for the octa BDE formulation, was 

0.002 ng/g wwt, and the mean total concentration over all 24 fish samples was 1.12 ng/g wwt. 

Schecter et al. (2009b) provided an update to this data by conducting a similar market 

basket survey in Texas in 2009. They obtained 310 samples of 31 food types, made composites 

of 10 for each food type, and analyzed the 31 final samples for 24 BDE congeners as well as 

hexabromocyclododecane (HBCD). Seven of these 31 samples were of fish, ranging from 

canned tuna to cod to frozen fish sticks. There was a somewhat similar issue relating to 

detection limits for this data set as was found for the earlier data set, in that the detection limit for 

some congeners approached the detected concentrations, particularly for BDE 209. Still, a 

decline in concentrations might be seen with this data. For example, the average concentrations 

of these four congeners (which were the congeners having the highest concentrations in 2006) 

were: BDE 47 at 0.2 ng/g wwt (it was 0.6 ng/g wwt in the earlier survey as noted above), BDE 

99 at 0.04 ng/g wwt (compared to 0.17 ng/g wwt), BDE 100 at 0.04 (compared to 0.13 ng/g 

wwt), and BDE 209 at 0.008 ng/g wt (calculated assuming ND = 0 for 5 nondetects; compared to 

0.098 ng/g wwt). As in 2006, BDE 153 and BDE 154 were detected frequently: 23 of 24 in 

2006, and 5 of 7 in 2009. Declines were seen in these congeners as well. BDE 153 declined 

from 0.02 to 0.006 ng/g wwt, and BDE 154 declined from 0.05 to 0.015 ng/g wwt. 

 Fish, meat, and fowl products were purchased in December 2003 and February 2004 

from three different food markets in Sacramento and El Dorado Hills in Northern California 

(Luksemburg et al., 2004). A total of 31 different BDE congeners were measured, although no 

congener-specific data were provided. Homologue BDE groups were measured, and tables were 

provided with data for those. The total concentrations found in fish ranged from 0.09 to 

4.9 ng/g wwt. 

 The Norwegian Institute of Public Health in Oslo, Norway conducted an intercomparison 

laboratory study on the measurement of PBTs including PCDD/Fs, PCBs, and PBDEs (Haug et 

al., 2005). A total of 73 laboratories participated, but only 21 laboratories reported back 

concentrations of seven PBDEs in samples of chicken, trout, and palm oil. The eight BDEs were 

BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209. Lake trout 

had the highest concentrations by a large margin: mean concentrations were 0.6; 95.2; 78.6; 39.2; 

9.9; 12.0; 0.02; and 0.06 ng/g wwt for the eight congeners, respectively, totaling 236 ng/g wwt. 

These results appear anomalous, not only because the concentration were so much higher than 
4-75

anything else in the literature (total PBDEs have been found in the single digit ng/g wwt or less, 

not in the hundreds of ng/g wwt), but also because their reported concentrations in chicken and 

palm oils were up to 4 orders of magnitude lower, at concentrations less than 0.01 ng/g wwt. 

Tittlemeier et al. (2004) purchased 122 fish and shellfish from retail stores in three 

Canadian cities (Vancouver, Halifax, and Toronto) in the winter of 2002 and analyzed them for 

18 BDE congeners. The fish types include salmon, trout, tilapia, Arctic char, mussels, oysters, 

shrimp, and crab, and the congeners include BDE 15, BDE 17, BDE 28, BDE 47, BDE 66, 

BDE 71, BDE 75, BDE 77, BDE 85, BDE 99, BDE 100, BDE 119, BDE 126, BDE 138, 

BDE 153, BDE 154, BDE 183, and BDE 190. Total BDEs per fish group ranged from 0.02 in 

shrimp to 1.6 ng/g wwt in trout. Salmon was second, with a geometric mean concentration of 

1.5 ng/g wwt. Trout and salmon had the highest lipid contents at 8% and 11%, respectively, 

partially explaining the high amount of BDEs in them. The third highest lipid content, at 7.9%, 

was found in Arctic char, and not unexpectedly, Arctic char had the third highest total BDE, at 

0.6 ng/g wwt. BDE 47 was the most predominant congener, explaining 48% of the 

concentration, with BDE 99 contributing 24%. Following these two were BDE 100, BDE 28, 

BDE 153, BDE 154, and BDE 183. Farmed samples showed higher concentrations than wild 

samples, particularly for salmon. 

Ohta et al. (2002) evaluated the concentration of PBDEs in breast milk and food 

products, including several fish species, in Japan. The congeners evaluated included BDE 28, 

BDE 47, BDE 99, BDE 153, and BDE 154. Concentrations in 19 fish samples (4 young 

yellowtail, 4 mackerel, 2 natural yellowtail, 3 salon, 3 yellow-fin tuna, and 3 short-necked clams) 

ranged between 0.02 and 1.65 ng/g wwt in the edible tissues of the fish, with the highest in 

yellow-fin tuna. The dominant congeners were BDE 47―at over 50% in the samples―and 

BDE 99―at about 20%. Questionnaires on food consumption were given to the women. In a 

“high” group of fish consumers (n = 5), the average breast milk concentration of BDEs was 

1,724 pg/g lwt. This was significantly higher than the average concentration of BDEs in a “low” 

group of fish consumers (n = 3), which was 774 pg/g lwt, suggesting that breast milk 

concentrations of BDEs were related to fish consumption. 

Pirard et al. (2005) presented results from salmon, whole trout, and Spanish mussels 

purchased from a Belgian supermarket. Total concentrations were 8.19, 2.69, and 10.22 ng/g lwt 
4-76

for mussel, trout, and salmon, respectively. BDE 47 dominated the results, comprising over 60% 

for the three samples, followed by BDE 99 (in mussel and trout) or BDE 100 (in salmon). 

 Meng et al. (2007) collected samples of 13 fish species from local fish markets and 

supermarkets from 11 fishery-producing regions in Guangdong Province, China. These include 

freshwater-farmed fish, seawater-farmed fish, and wild marine fish. Eleven congeners were 

measured, including BDE 209. The median and mean of 10 BDEs, not including BDE 209, were 

0.16 and 0.23 ng/g wwt, respectively. BDE 209 was found in only 14 of the 390 fish samples, 

ranging from <0.1 to 0.57 ng/g wwt, although the detection limit of BDE 209 was the highest, at 

0.1 ng/g wwt, compared to 0.001−0.003 ng/g wwt detection limits of other congeners. 

4.6.4. Observations from Fish Data

While this review falls short of a comprehensive review of the literature on brominated 

flame retardants in fish, it covers critical studies on fish from open aquatic settings (rivers and 

lakes), as well as from retail markets and fish farms. Measurements from open aquatic settings 

provide information on the understanding of environmental fate for this class of compounds, 

while sampling from retail markets or fish farms provide more direct information on exposure. 

Unfortunately, most of this sampling did not include the higher-brominated marker congeners 

BDE 183 and BDE 209. The very limited sampling for BDE 183 suggested insignificant 

concentrations, but the market basket surveys by Schecter et al. (2006a; 2009b) and the marine 

environment sampling by Johnson-Restrepo et al. (2005) suggested that BDE 209 can be 

significant and sometimes the highest congener found in fish. Most often, however, BDE 47 

dominated the profile, explaining over 50% of the concentration found, with BDE 99 the second 

highest found, explaining around 25%. The ratio of BDE 47 to BDE 99 was about 2.0 in these 

studies. Generally, total BDE concentrations were highest in open water environments (lakes, 

rivers, oceans) in contrast to farmed fish or fish obtained from marketplaces. In the wild 

environment, concentrations ranged above 1,000 ng/g wwt but most often were well above 

10 ng/g wwt, averaging between 10 and 100 ng/g wwt. Concentrations in farmed and market 

fish were generally lower, in the neighborhood of 1−5 ng/g wwt basis. It is not clear why 

store-bought fish might be lower than wild-caught fish, except possibly that the focus of 

wild-caught fish is from locations historically known to be impacted by contaminants such as 

dioxins, PCBs, or PBDEs, such as the Great Lakes. Like sediment cores (described in Section 
4-77

3.8) and body burden measurements (described in Section 5.1), temporal sampling of fish has 

provided evidence of the rise of these compounds in environmental matrices throughout the 

1990s into the 21st century. 

4.7. FOOD CONCENTRATIONS

 U.S. data on food other than fish are highlighted by four market basket surveys: two in 

Texas (Schecter et al., 2006a, 2009b), one in California (Luksemburg et al., 2004), and one 

sampling from several states in the United States (Huwe and Larsen, 2005). Other data include 

earlier sampling by Huwe et al. (2003) on chicken in a research mode and several studies from 

abroad. Similar to the fish data, there were inconsistent reports in terms of lipid or wwt. 

Table 4-4 provides congener-specific data on food concentrations. 

Schecter et al. (2006a; derived from data in Schecter et al., 2004, with additional data and 

analyses) reported on a sampling of 62 food items, split between samples of meat, dairy, and 

fish. BDE 17, BDE 28, BDE 47, BDE 66, BDE 77, BDE 85, BDE 99, BDE 100, BDE 138, 

BDE 153, BDE 154, BDE 183, and BDE 209 were detected, and, as described in Section 4.6.3, 

there may have been some analytical issues, given the high detection limits and a high frequency 

of nondetects in this data set. In addition to these limitations, all the food originates from Texas 

supermarkets only. Section 4.6.3 above describes the results for the 24 fish samples. Of 18 meat 

samples, total BDE ranged from 0.04 to 1.4 ng/g wwt. BDE 209 was detected in eight meat 

samples, with concentrations ranging from 0.01 to 0.25 ng/g wwt. Generally, the highest 

congeners were BDE 47 and BDE 99, with BDE 99 the highest in meat, averaging 

0.16 ng/g wwt; BDE 47 was at 0.09 ng/g wwt in meat. These two BDE congeners were similar 

in 15 dairy samples, averaging about 0.03 ng/g wwt. The mean of BDE 209 in dairy was 

0.04 ng/g wwt, but this was driven by a single sample of cream cheese at 0.48 ng/g wwt. All 

other samples were either nondetect (8 samples) or had concentrations under 0.02 ng/g wwt. The 

total average concentrations of BDE 209 in meat and dairy were 0.38 and 0.12 ng/g wwt, 

respectively. Two samples of infant formula were included in this survey, and these were the 

only infant formula data that could be found. The wwt concentrations in the samples were 

32 and 25 pg/g wwt, respectively, and BDE 209 made up about half of the concentration. No 

data on solid baby foods could be found in the literature. 
4-78

Table 4-4. Congener-specific concentrations of PBDEs in food originating from the 

United States (concentrations in ng/g wwt or specifically identified as ng/g lwt, if 

that was how reported by the author and conversion to ng/g wwt was not possible)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

TriBDE

17 0.0008 n = 18 meat (lipid = 26.3%) including pork, chicken, beef; 

dominated by NDs at DL = 0.7

Schecter et al. 

(2006a)

0.0003 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream

Schecter et al. 

(2006a)

ND n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef, DL < 0.0002

Schecter et al. 

(2009b)

ND n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, DL < 0.001

Schecter et al. 

(2009b)

<0.00001 n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), one detect for apple at 0.00004 

Schecter et al. 

(2009b)

ND n = 1 egg (lipid = 10%), DL = 0.0001 Schecter et al. 

(2009b)

17/25 0.008 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003) 

0.008 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003)

28 0.005 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al. 

(2006a)

0.0008 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream

Schecter et al. 

(2006a)

0.0006 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; one ND

Schecter et al. 

(2009b)

0.0034 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt

Schecter et al. 

(2009b)

0.0009 n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), two detects including cereal at 

0.003 

Schecter et al. 

(2009b)

ND n = 1 egg (lipid = 10%), DL = 0.00009 Schecter et al. 

(2009b)

33 0.017 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003) 

0.028 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003)

28/33 0.004 (lwt) Market basket hamburger (n = 11), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

0.004 (lwt) Market basket bacon (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)
4-79

Table 4-4. Congener specific concentrations of PBDEs in food originating 

from the United States (concentrations in ng/g wwt or specifically identified as 

ng/g lwt, if that was how reported by the author and conversion to ng/g wwt 

was not possible) (continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.003 (lwt) Market basket chicken fat (n = 22), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.007 (lwt) Market basket pork fat (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.002 (lwt) Market basket beef fat (n = 10), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

TetraBDE

47 2.26 9 chicken fat (90% lipid) samples from 3 production 

noncontaminated sites

Huwe et al.

(2003)

6.87 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003) 

0.093 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al. 

(2006a)

0.032 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream

Schecter et al. 

(2006a)

0.016 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef

Schecter et al. 

(2009b)

0.134 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt

Schecter et al. 

(2009b)

0.007 n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit) 

Schecter et al. 

(2009b)

0.009 n = 1 egg (lipid = 10%) Schecter et al. 

(2009b)

0.18 (lwt) Market basket hamburger (n = 11), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

0.23 (lwt) Market basket bacon (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.81 (lwt) Market basket chicken fat (n = 22), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

1.07 (lwt) Market basket pork fat (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.07 (lwt) Market basket beef fat (n = 10), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

49 ND n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; DL < 0.0004

Schecter et al. 

(2009b)
4-80

Table 4-4. Congener specific concentrations of PBDEs in food originating 

from the United States (concentrations in ng/g wwt or specifically identified as 

ng/g lwt, if that was how reported by the author and conversion to ng/g wwt 

was not possible) (continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

ND n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, DL < 0.0005 (DL for butter at 0.003)

Schecter et al. 

(2009b)

0.00003 n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), 1 detect for apple at 0.0002 

Schecter et al. 

(2009b)

ND n = 1 egg (lipid = 10%), DL = 0.0002 Schecter et al. 

(2009b)

66 0.001 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003)

0.018 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003) 

0.0001 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al. 

(2006a)

0.0006 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream

Schecter et al. 

(2006a)

ND n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; DL < 0.0004

Schecter et al. 

(2009b)

ND n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, DL < 0.0005 (DL for butter at 0.003)

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit); DL < 0.003 

Schecter et al. 

(2009b)

ND n = 1 egg (lipid = 10%), DL = 0.0002 Schecter et al. 

(2009b)

77 0.0008 n = 18 meat (lipid = 26.3%) including pork, chicken, beef; 

mostly ND or NA

Schecter et al. 

(2006a)

0.0001 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream; all ND or NA

Schecter et al. 

(2006a)

PentaBDE

85 0.11 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003)

0.52 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003)

0.005 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al. 

(2006a)

0.001 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream; most ND or NA except one 

cheese at 5.52

Schecter et al. 

(2006a)
4-81

Table 4-4. Congener specific concentrations of PBDEs in food originating 

from the United States (concentrations in ng/g wwt or specifically identified as 

ng/g lwt, if that was how reported by the author and conversion to ng/g wwt 

was not possible) (continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.022 (lwt) Market basket hamburger (n = 11), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

0.023 (lwt) Market basket bacon (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.047 (lwt) Market basket chicken fat (n = 22), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.034 (lwt) Market basket pork fat (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.014 (lwt) Market basket beef fat (n = 10), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

0.0004 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; 1 detect; DL < 0.0006

Schecter et al. 

(2009b)

0.0007 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, 1 detect; DL <0.001 except butter at 0.005

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.008 

Schecter et al. 

(2009b)

ND n = 1 egg (lipid = 10%), DL = 0.0004 Schecter et al. 

(2009b)

99 3.11 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003)

10.87 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003)

0.16 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al. 

(2006a)

0.028 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant

formula, yogurt, ice cream

Schecter et al. 

(2006a)

0.26 (lwt) Market basket hamburger (n = 11), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

0.018 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef

Schecter et al. 

(2009b)

0.098 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt

Schecter et al. 

(2009b)

0.0076 n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit)

Schecter et al. 

(2009b)

0.019 n = 1 egg (lipid = 10%) Schecter et al. 

(2009b)
4-82

Table 4-4. Congener specific concentrations of PBDEs in food originating 

from the United States (concentrations in ng/g wwt or specifically identified as 

ng/g lwt, if that was how reported by the author and conversion to ng/g wwt 

was not possible) (continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.30 (lwt) Market basket bacon (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

1.38 (lwt) Market basket chicken fat (n = 22), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

1.04 (lwt) Market basket pork fat (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.11 (lwt) Market basket beef fat (n = 10), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

100 0.45 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003) 

1.47 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003)

0.023 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al. 

(2006a)

0.005 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream

Schecter et al. 

(2006a)

0.003 n=8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef

Schecter et al. 

(2009b)

0.016 (lwt) n=8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt

Schecter et al. 

(2009b)

0.0001 n=7 vegetable based (lipid = 50 -100% for oil/peanut 

butter; 0 for cereals/fruit), one detect for apple at 0.0009 

ng/g wwt 

Schecter et al. 

(2009b)

0.007 n=1 egg (lipid = 10%) Schecter et al. 

(2009b)

0.042 (lwt) Market basket hamburger (n = 11), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

0.041 (lwt) Market basket bacon (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.28 (lwt) Market basket chicken fat (n = 22), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.18 (lwt) Market basket pork fat (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.02 (lwt) Market basket beef fat (n = 10), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

119 ND n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; DL < 0.0006

Schecter et al. 

(2009b)
4-83

Table 4-4. Congener specific concentrations of PBDEs in food originating 

from the United States (concentrations in ng/g wwt or specifically identified as 

ng/g lwt, if that was how reported by the author and conversion to ng/g wwt 

was not possible) (continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

ND n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, DL < 0.001 except for butter at 0.004

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.006 

Schecter et al. 

(2009b)

ND n = 1 egg (lipid = 10%), DL = 0.0004 Schecter et al. 

(2009b)

126 ND n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; DL < 0.0004

Schecter et al. 

(2009b)

HexaBDE

138 0.02 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003)

0.16 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003) 

0.002 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al. 

(2006)

0.0003 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream; most ND or NA with high DL

Schecter et al. 

(2006a)

0.0005 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; 2 detects; DL < 0.0006

Schecter et al. 

(2009b)

ND n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, DL < 0.001 except for butter at 0.006

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.006 

Schecter et al. 

(2009b)

ND n = 1 egg (lipid = 10%), DL = 0.0004 Schecter et al. 

(2009b)

140 0.003 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003) 

0.015 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003)

153 060 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003)

2.90 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003) 

0.021 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al., 

(2006a)

0.004 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream

Schecter et al. 

(2006a)
4-84

Table 4-4. Congener specific concentrations of PBDEs in food originating 

from the United States (concentrations in ng/g wwt or specifically identified as 

ng/g lwt, if that was how reported by the author and conversion to ng/g wwt 

was not possible) (continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.004 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef

Schecter et al. 

(2009b)

0.0128 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt

Schecter et al. 

(2009b)

0.00006 n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), one detect for apple at 

0.0004 ng/g wwt 

Schecter et al. 

(2009b)

0.007 n = 1 egg (lipid = 10%) Schecter et al. 

(2009b)

0.10 (lwt) Market basket hamburger (n = 11), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

0.078 (lwt) Market basket bacon (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.22 (lwt) Market basket chicken fat (n = 22), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.12 (lwt) Market basket pork fat (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.019 (lwt) Market basket beef fat (n = 10), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen, (2005)

154 0.18 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003) 

0.55 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003)

0.014 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al. 

(2006a)

0.002 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream

Schecter et al. 

(2006a)

0.002 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef

Schecter et al. 

(2009b)

0.0048 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.007 

Schecter et al. 

(2009b)

0.004 n = 1 egg (lipid = 10%), DL = 0.00011 Schecter et al. 

(2009b)

0.029 (lwt )Market basket hamburger (n = 11), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)
4-85

Table 4-4. Congener specific concentrations of PBDEs in food originating 

from the United States (concentrations in ng/g wwt or specifically identified as 

ng/g lwt, if that was how reported by the author and conversion to ng/g wwt 

was not possible) (continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.042 (lwt) Market basket bacon (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.088 (lwt) Market basket chicken fat (n = 22), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.087 (lwt) Market basket pork fat (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.011 (lwt) Market basket beef fat (n = 10), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

156 ND n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; DL < 0.001

Schecter et al. 

(2009b)

HeptaBDE

183 0.19 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003) 

0.34 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003)

0.01 n = 18 meat (lipid = 26.3%) including pork, chicken, beef Schecter et al. 

(2006a)

0.002 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream

Schecter et al. 

(2006a)

0.008 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; one ND

Schecter et al. 

(2009b)

0.0103 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt; 3 detects; butter DL at 0.008

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.012 

Schecter et al. 

(2009b)

ND n = 1 egg (lipid = 10%), DL = 0.0008 Schecter et al. 

(2009b)

0.029 (lwt) Market basket hamburger (n = 11), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)

0.12 (lwt) Market basket bacon (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.13 (lwt) Market basket chicken fat (n = 22), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.084 (lwt) Market basket pork fat (n = 11), 10 locations around the 

country, sampled in 2005

Huwe and 

Larsen (2005)

0.012 (lwt) Market basket beef fat (n = 10), 10 locations around the 

country, sampled in 2004

Huwe and 

Larsen (2005)
4-86

Table 4-4. Congener specific concentrations of PBDEs in food originating 

from the United States (concentrations in ng/g wwt or specifically identified as 

ng/g lwt, if that was how reported by the author and conversion to ng/g wwt 

was not possible) (continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

196 0.002 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; 2 detects

Schecter et al. 

(2009b)

0.002 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, 1 detect butter at 0.014 

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.007 

Schecter et al. 

(2009b)

0.0006 n = 1 egg (lipid = 10%) Schecter et al. 

(2009b)

197 0.004 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; 2 detects

Schecter et al. 

(2009b)

0.002 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, 1 detect butter at 0.011

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.007 

Schecter et al. 

(2009b)

ND n = 1 egg (lipid = 10%), DL = 0.0004 Schecter et al. 

(2009b)

OctaBDE

206 0.001 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; 2 detects

Schecter et al. 

(2009b)

0.037 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, 2 detects including butter at 0.224

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.013 

Schecter et al. 

(2009b)

0.005 n = 1 egg (lipid = 10%) Schecter et al. 

(2009b)

207 0.003 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; one ND

Schecter et al. 

(2009b)

0.059 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, 2 detects including butter at 0.359

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.013 

Schecter et al. 

(2009b)

0.003 n = 1 egg (lipid = 10%) Schecter et al. 

(2009b)

DecaBDE

209 1.24 9 chicken fat (90% lipid) samples from three production 

noncontaminated sites

Huwe et al.

(2003) 
4-87

Table 4-4. Congener specific concentrations of PBDEs in food originating 

from the United States (concentrations in ng/g wwt or specifically identified as 

ng/g lwt, if that was how reported by the author and conversion to ng/g wwt 

was not possible) (continued)

Congener Concentration

(ng/g wwt or lwt) Comment Reference

0.72 4 chicken fat (90% lipid) samples from a contaminated 

production site

Huwe et al.

(2003)

0.053 n = 18 meat (lipid = 26.3%) including pork, chicken, beef; 

dominated by 1 turkey at 245

Schecter et al. 

(2006a)

0.041 n = 15 dairy (lipid = 10.2%) including cheese, milk, infant 

formula, yogurt, ice cream; dominated by 1 sample at 481; 

others at <20

Schecter et al. 

(2006a)

0.017 n = 8 meat (lipid = 13.3%) including pork, chicken, turkey, 

beef; 5 detects

Schecter et al. 

(2009b)

0.954 (lwt) n = 8 dairy (lipid = 26%) including butter, cheese, milk, ice 

cream, yogurt, 6 detects including butter at 5.19

Schecter et al. 

(2009b)

ND n = 7 vegetable based (lipid = 50−100% for oil/peanut 

butter; 0 for cereals/fruit), DL < 0.08 

Schecter et al. 

(2009b)

0.034 n = 1 egg (lipid = 10%), DL = 0.00011 Schecter et al. 

(2009b)

Schecter et al. (2009b) provided an update to this data by conducting a similar market 

basket survey in Texas in 2009. They obtained 310 samples of 31 food types, made composites 

of 10 for each food type, and analyzed the 31 final samples for 24 BDE congeners as well as 

hexabromocyclododecane. Twenty-four of these 31 samples were of meat, dairy, and products 

of vegetable origin. Declines were seen in most food types. Total BDEs (sum of 20 congeners 

with sufficient detections, including BDE 196, BDE 197, BDE 206, BDE 207, and BDE 209) for 

meats ranged from 0.008 to 0.2 ng/g wwt, with BDE 209 being found at no higher than 0.04 ng/g 

wwt. Dairy concentrations of total PBDEs in seven of eight samples were lower than 0.3 ng/g 

wwt, but was 6.2 ng/g wwt in a sample of butter (90% lipid) with BDE 209 at 5.2 ng/g wwt. As 

in the previous survey, BDE 47 tended to dominate the concentrations, but with significant 

contributions also from BDE 99, BDE 100, BDE 153, and BDE 154. BDE 209 was detected in 

five of eight meat samples ranging from 0.009 to 0.04 ng/g wwt, and in six of eight dairy 

samples ranging from 0.004 to 5.2 ng/g wwt. It was the highest congener found in eggs at 

0.03 ng/g wwt. 
4-88

 Fish, meat, and fowl products were purchased in December 2003 and February 2004 

from three different food markets in Sacramento and El Dorado Hills in Northern California 

(Luksemburg et al., 2004). Thirty-one different BDE congeners were measured, although no 

congener-specific data were provided. Homologue BDE groups were measured and tables were 

provided with data for those. The total PBDE concentrations (sum of homologue concentrations 

provided) were higher in fish (from 0.09 to 4.9 ng/g wwt) and fowl (0.09 to 2.5 ng/g wwt) than 

in beef and deer meat products (from 0.1 to 0.4 ng/g wwt). It was stated that the highest 

concentrations of individual congeners were BDE 47, BDE 99, and BDE 100, although 

decaBDE, which is the single congener BDE 209, most often was the highest homologue group 

found. 

A nonstatistical sampling of market basket meat and poultry items from large 

supermarkets was undertaken in 2004, in these states: FL, VA, CT, PA, ND, MT, OR, NM, and 

AZ (Huwe and Larsen, 2005). A total of 65 meat samples, including hamburger (n = 11), bacon 

(n = 11), chicken fat (n = 22), pork fat (n = 11), and beef fat (n = 10), were measured for 

BDE 28/33, BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183. Beef had 

the lowest average concentrations of BDEs (0.25 ng/g lwt for beef fat and 0.67 ng/g lwt for 

hamburger), while chicken and pork had the highest average concentrations―2.96 ng/g lwt and 

2.62 ng/g lwt, respectively. These average concentrations were elevated somewhat by a few 

high samples: One pork and two chicken samples had total BDEs greater than 15.0 ng/g lwt. 

BDE 99 dominated all food types, thus, accounting for 39%, 36%, 46%, 40%, and 44% for 

hamburger, bacon, chicken fat, pork fat, and beef fat respectively. BDE 47 was second most 

dominant, explaining 27%, 27%, 27%, 41%, and 28%, respectively. The ratio of BDE 47 to 

BDE 99 averaged 0.78, which is similar to the BDE 47 to BDE 99 ratio of 0.6 found in a penta 

formulation. This contrasts with a BDE 47 to BDE 99 ratio of sometimes greater than 2, which 

was identified in several studies in fish as described in Section 4.6. As will be described in the 

next chapter, humans also have more BDE 47 congeners than BDE 99, but about this same 

factor of 2.0. 

Huwe et al. (2003) earlier conducted research-oriented studies on BDEs in poultry. 

Chicken fat samples from three chicken production sites known to have chickens “contaminated” 

by consumption of animal feed with ball clay, which had high levels of dioxins, and two other 

chicken fat samples from an uncontaminated production site were analyzed for BDEs. The 
4-89

congeners analyzed included the following: BDE 17/25, BDE 33, BDE 47, BDE 66, BDE 85, 

BDE 99, BDE 100, BDE 138, BDE 140, BDE 153, BDE 154, BDE 183, and BDE 209. Four 

other unidentified congeners were noted, but they are not included in the summary. The samples 

from two of the “contaminated” sites did not appear to have BDE levels different from the 

noncontaminated production site. The nine samples from these three locations were averaged, 

and the four samples from the contaminated site, where levels were higher, were analyzed 

separately. The authors conducted a PBDE analysis of the stored contaminated feed but did not 

find elevated levels of PBDEs (data not provided). This suggested to them that feed was not the 

cause of high levels of PBDEs in the one site. The authors did state that in the production site 

where BDE levels were higher, a factory producing pentaBDE formulations was located in the 

same city, which may have explained the higher levels found there. The average total BDE 

concentration of nine “uncontaminated” chicken samples was 8.3 ng/g lwt, and the average of 

four “contaminated” chicken samples was 24.8 ng/g lwt. BDE 99 was the highest congener; it 

averages 3.1 ng/g lwt uncontaminated and 10.9 ng/g lwt contaminated. BDE 47 averaged 

2.3 ng/g lwt uncontaminated and 6.9 ng/g lwt contaminated. BDE 209 was generally low in all 

samples, with actually lower concentrations in the contaminated versus uncontaminated sites; the 

average over all 13 samples was 1.1 ng/g lwt. 

 Studies of PBDEs in food abroad mostly originate from Europe, and results are 

comparable to America in that total PBDEs tend to be less than 1 ng/g, dominated by BDE 47 

and BDE 99, with sparse data on BDE 209. As noted below, however, three studies have 

documented high levels of BDE 209 in food samples. In the United Kingdom study, BDE 209 

was found at the highest concentration of all congeners in nearly every sample; in the study from 

Spain, BDE 209 was found at the highest level in several food products; and in the study from 

Japan, BDE 209 dominated oils and fats. In all three studies, high concentrations of BDE 209 

were 10 times or more higher than other congeners when BDE 209 was found. 

 A recent study commissioned by WWF (formerly World Wildlife Fund, now just WWF), 

conducted by the Netherlands Organization for Applied Scientific Research (TNO) (Peters, 

2006), measured 30 BDEs in 26 food products from different countries in Europe. The food 

products ranged from honey in the United Kingdom, to salami in Italy, to pork chops in Poland. 

BDEs were found in 19 of 26 products: a low total of 0.15 ng/g found in honey to a high total of 

1.3 ng/g wwt in minced beef. BDE 209 was tested for but not found, although it cannot be 
4-90

expected to be found with a high detection limit of 5.0 ng/g wwt. Like other studies, BDEs 47 

was found most frequently; it was found in 17 of 26 samples at an average positive concentration

of 0.49 ng/g. Interestingly, BDE 32, not measured in other studies, was found next most 

frequently, in 12 of 26 samples, with an average positive concentration of 0.07 ng/g wwt. 

BDE 99 was found in 8 of 26 samples at an average positive concentration of 0.12 ng/g. 

 Another European study included a market basket survey of BDEs, including BDE 209, 

from Belgium (Voorspoels et al., 2007), although BDE 209 was not quantified in any sample. 

Concentrations generally were all less than 0.1 ng/g wwt, with fish being the highest generally, 

and only 2 of 7 food samples contained BDE concentrations greater than 1.0 ng/g wwt, at 1.0 and 

1.6 ng/g wwt. Butter was the highest of the nonfish samples, at 0.8 ng/g wwt. The eight meat 

samples were all under 0.2 ng/g wwt, as were fast food and eggs. Cheese was slightly higher at 

0.22 ng/g wwt.

 Three recent surveys showed high levels of BDE 209, one from United Kingdom (FSA, 

2006), one from Spain (Gomara et al., 2006), and one from Japan (Akutsu et al., 2008). In the 

dietary survey conducted by the Food Standards Agency of the United Kingdom, composite 

samples representing 19 food groups were analyzed for a suite of 17 BDE congeners, including

BDE 183 and BDE 209. The meat product concentration of BDE 209 was exceedingly high at 

3.64 ng/g wwt. In fact, BDE 209 had the highest concentration of all congeners in all but two

composite food group samples, although concentrations other than the meat concentration appear 

more in line with other values in the literature, at less than 0.5 ng/g wwt. For example, BDE 209 

was found at 0.29 in “fats and oils,” and the next highest congener concentrations were BDE 49 

and BDE 99, both reported at 0.08 ng/g wwt. Although summary statistics were not supplied, 

generally BDE 47 and BDE 99 were similar, at concentrations between 0.01 and 0.10 ng/g, 

consistent with other surveys, and other congeners were present but at lower concentrations. 

Gomara et al. (2006) collected 104 Spanish food samples randomly from local supermarkets all 

over Spain from 2003 to 2005. The samples encompassed 21 types of food, including milk and 

dairy products, eggs, fish (tuna, sardine, and others), meat and meat products, vegetable oil, and 

shellfish. A total of 15 BDEs were measured, including for the first time the higher-brominated 

BDE 184, BDE 191, BDE 196, and BDE 197, in addition to BDE 183 and BDE 209. The 

highest total median concentration was found in fish, 189 pg/g wwt, followed by oils at 

119 pg/g wwt, meats at 76 pg/g wwt, shellfish at 76 pg/g wwt, eggs at 74 pg/g wwt, and dairy at 
4-91

66 pg/g wwt. BDE 209 was the predominant congener in oil and egg samples, at 25 and 

37 pg/g wwt, respectively, but it was also found at significant levels in all other food products, 

with medians in dairy products at 4 pg/g wwt, in meats at 11 pg/g wwt, in fish at 5 pg/g wwt, and 

shellfish at 7 pg/g wwt. Otherwise, BDE 47 and BDE 99 were the predominant congeners in 

dairy products (11 and 9 pg/g wwt, respectively), meats (17 and 15 pg/g wwt), fish (115 and 

15 pg/g wwt), and shellfish (11 and 5 pg/g wwt). The study from Japan was a total dietary 

survey, like the UK survey, and entailed measurements of 36 BDE congeners in 14 food groups. 

Detectable levels of one or more congeners were found in four food groups. Total 

concentrations (sum of 36 congeners) in these were 1.8 ng/g wwt in “fats and oils” dominated by 

BDE 209 at 1.8 ng/g lwt, with detectable levels (0.01 ng/g wwt) of BDE 28/33 and BDE 47, 

“legumes and their products” at a low level of 0.03 ng/g wwt, “fish, shellfish, and their products” 

at a level of 0.48 ng/g wwt, dominated by BDE 47 at 0.22 ng/g wwt with detections of eight 

other congeners, and “meat and eggs” at a low level of 0.01 ng/g wwt. Eighteen samples of 

vegetable oil were resampled to verify the high BDE 209 finding, and ng/g levels of BDE 209 

(0.7−2.4 ng/g wwt) were found in seven of them. 

 Earlier in 2003, Bocio et al. (2003) evaluated dietary exposure of individuals in Spain to 

PBDEs. A total of 54 samples were developed as composites of numerous food types, including 

vegetables, tubers, pulses (peas, beans, and lentils), cereals, fruits, fish, and shellfish, meat and 

meat products (pork, chicken, beef, lamb) eggs, milk, dairy, fats and oils. Samples were 

analyzed for total homologue groups, so no individual congeners were measured. Because of 

this, the “total” concentrations here probably should not be compared with totals from other 

studies that measured and reported on individual congeners. The highest concentration of total 

PBDEs was found in oils and fats (0.6 ng/g wwt), followed by fish and shellfish (0.3 ng/g wwt), 

meat and meat products (0.1 ng/g wwt), and eggs (0.06 ng/g wwt), with essentially none for 

vegetables and grains. A predominance of the tetra and penta homologues, followed by hexa 

congeners, was found in the samples. 

 As described in the section above on fish, the Norwegian Institute of Public Health in 

Oslo, Norway conducted an intercomparison laboratory study on the measurement of PBTs, 

including PCDD/Fs, PCBs, and PBDEs (Haug et al., 2005). It was noted that fish concentrations 

were unusually high, totaling 236 ng/g wwt. Results for the other congeners: BDE 28, BDE 47, 

BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209, were more in line with other 
4-92

values from around with world. Levels of these eight congeners in chicken were 0.0007, 0.02, 

0.02, 0.007, 0.005, 0.003, 0.003, and 0.07 ng/g wwt, thus totaling 0.14 ng/g wwt, and palm oil 

levels were 0.005, 0.02, 0.04, 0.02, 0.008, 0.006, 0.007, and 0.39 ng/g wwt, thus totaling

0.49 ng/g wwt. BDE 209 was the highest in these chicken and palm oil samples, while it was 

among the lowest in the fish samples reported earlier. 

 Irish animal and vegetative food products were surveyed by Tlustos et al. (2005) for the 

presence of PBDEs. A total of 65 samples, most of them representing pooled samples of 10 or 

more, were analyzed for BDE 17, BDE 28, BDE 47, BDE 47, BDE 66, BDE 71, BDE 77, 

BDE 85, BDE 99, BDE 100, BDE 119, BDE 126, BDE 138, BDE 153, BDE 154, and BDE 183. 

Only total concentrations were reported, with no discussion on congener distribution in the food 

products. Food products included carcass fat of bovine, avian (including duck), ovine, and 

porcine; dairy products included butter, cheddar cheese, soft cheese, processed cheese, dairy 

spread, and yogurt; liver of bovine, avian (chicken and turkey), ovine, and porcine; soup; cereals; 

fruit; vegetables; vegetable/animal fat; and vegetable oil. Concentrations of total PBDEs ranged 

narrowly from 0.85 to 1.49 ng/g lwt in all the terrestrial animal food products, and 

0.17−0.34 ng/g wwt in the cereals, fruit, vegetables, and vegetable oil. 

 Levels of 47, 99, 100, 153, and 154 were measured in 33 food items between 2002 and 

2004 in Norway (Knutsen et al., 2005). Levels ranged from 0.1−0.5 ng/g lwt for most terrestrial 

animal food products, but there were high measurements in eggs at 3.85 ng/g lwt, margarine 

(with 5% fish oil) at 3.08 ng/g lwt, liver paste at 1.17 ng/g lwt, and pork liver at 1.08 ng/g lwt. 

Levels were much higher in fish, with wwt concentrations for the majority of fish types above 

1.0 ng/g wwt, with a high of 9.73 ng/g wwt for cod liver oil. 

Harrad et al. (2004) presented data on BDE 47, BDE 99, BDE 100, BDE 153, and 

BDE 154 for indoor and outdoor air, and meat and vegan diets. The purpose of these 

measurements was to estimate daily exposure to BDEs via inhalation and diet. The total PBDE 

concentration in the composited samples of vegan and omnivorous diets were 0.15 and 

0.18 ng/g dwt, respectively. Similar to air, BDE 47, BDE 99, and BDE 100 were higher in meat 

diets as compared to vegan diets―BDE 153 and BDE 154 were very similar. Meat diet samples 

averaged 0.07, 0.07, and 0.02 ng/g dwt for BDE 47, BDE 99, and BDE 100, while the analogous 

concentrations for vegan samples were 0.05, 0.06, and 0.01 ng/g dwt. 
4-93

Schecter et al. (2006b) looked at the changes in PBDE levels when beef, lamb, and fish 

were cooked. While they reported a reduction in the amount of total BDEs in the samples, 

concentrations before and after cooking were not reported. They did show how the percent lipid 

was reduced by cooking (by fat dripping away), but concentrations are estimated as the mass of 

BDEs divided by the mass of lipids (or whole weight), and before and after masses were not 

provided. Still, the reported amount of total BDEs decreased by over 60% in beef and lamb, by 

over 50% in catfish, and by smaller amounts (~10%) in trout and salmon. This certainly 

suggests that concentrations also would have been reduced, and it would have been helpful if the 

authors had provided concentrations in their article. While their recommendation that exposures 

consider cooked foods is reasonable, most often exposure via food is calculated by 

concentrations provided in measurements from food products that are uncooked due to lack of 

information on concentrations in cooked foods. 

In summary, total BDE concentrations in terrestrial food products seem to be lower (less 

than 1 ng/g wwt) than fish BDE concentrations (1−5 ng/g wwt). As for fish, a paucity of data on 

BDE 209 makes it hard to generalize, although when sampled for, it appeared to be present at 

levels comparable to those for BDE 47 and BDE 99―the congeners generally found at the 

highest concentration. BDE 99 was typically found at the highest concentration, about twice as 

high as BDE 47. This contrasts with the relationship between these two congeners for fish, 

where BDE 47 is about twice as high as BDE 99. In very limited sampling, concentrations 

appear to be much lower in food products of vegetative origin (such as cereals, fruits, and 

vegetables) as compared to terrestrial animal food products. This is to be expected, as these 

organic PBTs tend to bioaccumulate in fat of animals. 

4.8. ASSIGNING EXPOSURE MEDIA CONCENTRATIONS FOR EXPOSURE 

ASSESSMENT PURPOSES

 The challenges with assigning values for individual congeners in exposure media are 

many: 

1. There is no consistent set of measured congeners, as in the case of the 17 toxic dioxin and 

furan congeners. While the California Air Resources Board studies in air entailed 

33 congeners (CARB, 2005), most studies only measured a handful of congeners. 

Moreover, the majority of past studies have not measured the deca congener, BDE 209, 

which remains the primary congener in currently produced and marketed products. The 
4-94

second congener for which limited data exist is BDE 183. The octa formulation of 

PBDEs contains about 44% hepta congeners, the most of any homologue group, and 

BDE 183 is the dominant stable congener within that homologue group. Its presence in 

an environmental matrix could only have occurred via debromination of 

higher-brominated congeners or because of the presence of the octa formulation. For this 

reason, BDE 183 is generally considered a primary marker for the possible presence of 

the octa BDE formulation. 

2. There are no statistically designed surveys that would provide a rigorous estimate of 

exposure media concentrations to which the general population of the United States is 

exposed. Most studies are nonstatistical targeted surveys, conducted within a small 

geographic area, and driven by a limited budget. 

3. Key exposure matrices remain sparsely studied, including drinking water, indoor air, 

outdoor soil, and animal food products of terrestrial origin (meat, dairy, and eggs). Food 

concentration data have only been obtained in the context of a limited number of retail 

market basket surveys. While these have merit, their coverage is limited: only one in 

three had any dairy samples, and the one study that did obtain samples from several states 

(Huwe and Larsen, 2005) did not include measurements for BDE 209. 

 Therefore, it is not possible to derive media concentrations that are statistically 

representative of general population exposures. Nonetheless, reasonable assumptions can be 

made for exposure media concentrations to use in making exposure estimates (see Chapter 5). 

Decisions needed to be made on which congeners to include in this derivation, and which studies 

to rely upon. The key congeners that have been measured include the primary congeners of the 

penta formulation, including BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154. BDE 183 

and BDE 209 remain of interest for their representation of the octa and deca formulations, as 

described above. The U.S. Environmental Protection Agency (EPA) PBDE project plan2

 

2

See http://www.epa.gov/oppt/pbde. 

identifies BDE 28 as a triBDE of interest, BDE 85 as a tetraBDE of interest, BDE 197 as an 

octaBDE of interest, and BDE 206 as nonaBDE of interest, so these will be included. Finally, 

many studies have also included triBDE 17, tetraBDE 66, and hexaBDE 85 in their 

measurements; so these final congeners will be added, bringing the total to 14 congeners for 

which profiles will be derived, if the data permit. There are limited concentrations available for 

other congeners, as displayed in the media-specific tables earlier in this chapter, but only these 

14 will be assigned values used in exposure calculations in the next chapter. 
4-95

Determining final concentrations to represent the general U.S. population exposure is not 

straightforward. Some key considerations for this compilation include the following: 

1. The studies should come from the United States, or perhaps Canada. Only when North 

American studies are unavailable will European or other foreign studies be used. 

2. Occupational data, while of interest, do not represent general population exposure. For 

indoor, as well as outdoor, exposures, it should be clear that the data were not taken in the 

vicinity of known sources, such as recycling facilities, autoshredding facilities, 

manufacturing facilities, and the like. 

3. Studies with a full suite of congeners, and, in particular, BDE 183 and BDE 209, are 

preferable to studies that have the limited set of BDEs associated with the penta 

formulation―BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154. 

4. Attempts to average congener-specific data across studies should be done with caution, if 

at all. For example, it may be preferable to use data from one geographic area, as long as 

it is appropriately background and not occupational, if those data have a full congener 

suite For example, relying on an entire set of urban air data taken by the California Air 

Resources Board, including the standard BDE 47, BDE 99, etc, as well as critical 

BDE 183 and BDE 209, would be preferable to only using the BDE 183 and BDE 209 

data from CARB, while averaging the CARB data on other congeners with data from 

other urban or rural locations in the United States. 

5. EPA has not provided an independent review of the quality of data contained in the 

studies that it has reviewed in this chapter, nor will EPA attempt to independently verify 

the quality of data in the studies it selects to represent exposure media concentrations. 

Rather, best judgment is used to select studies that contain high quality data. In addition 

to the factors cited above, additional considerations that speak to data quality include 

identification of a laboratory of known high quality, reputation of the authors of the 

study, and generally, the publication of the study in a respected peer-reviewed journal. 

 With that as backdrop, Table 4-5 contains the final derived profiles for water, surface 

soil, house dust, indoor and outdoor air, and categories of food products for use in the exposure 

calculations in the next chapter. When averages were required that were derived from individual 

sample data, it was assumed that ND = 0 for these calculations. This was done because often the 

listed detection limit, such as those from Schecter et al. (2009b), were much higher than the 

detected quantifications (see Section 4.6 for more detail and examples on the detection limit 

issue in the Schecter et al., 2009b effort). Following now is a brief justification for each of the 

media:
4-96

1. Drinking water: Of the three studies found for surface water (none were found for ground 

water) in the United States, the results from the San Francisco Estuary (Oros et al., 2005) 

were used to represent drinking water exposures. Concentrations of total PBDEs were 

similar in the three studies, near or below 100 pg/L, but the study of the San Francisco 

Estuary included BDE 209. However, the Estuary is not a drinking water source, and 

issues of turbidity and salinity could influence the results found. Since concentrations 

were similar in all studies available, this is not expected to present a problem for intake 

calculations; the more important issue is that drinking water data are not available. Total

PBDE concentrations ranged from 3 to 513 pg/L, with a mean of 146.2 pg/L. The mean 

concentration of BDE 209 in 18 measurements was 42.3 pg/L. 

2. Surface soil: The only systematic study of surface soil concentrations not associated with 

an industrial or other contaminated source for United States was from Offenberg et al. 

(2006), and results from this study will be used in this exercise. The average 

concentration over the 14 congeners of this assessment was 82 ng/g dry. The only other 

data on background soils came from Hassanin et al. (2004). It represented European soils 

and had a much lower total PBDE concentration of only about 1 ng/g dwt. 

3. House dust: The Environmental Working Group (Sharp and Lunder, 2004) data set 

includes samples provided by 10 women who earlier participated in a breast 

milk-sampling program. The concentrations of the congeners track well with other data, 

as seen in Table 4-2. Because these data originated from nine different states, it was 

judged that this might be the most representative data set. The average total PBDEs of 

8,275 ng/g dwt for the 14 congeners compares with an average of 5,811 ng total PBDEs/g 

dwt from 17 homes in the Washington, DC area (Stapleton et al., 2005); the 9,271 ng 

total PBDE/g dwt average from two samples from a computer lab in California (CARB, 

2005); and the geometric means of three locations (living room, bedroom, and from a 

household vacuum) within 20 homes in Boston of 13,732; 6,255; and 4,269 ng/g dwt, 

respectively (Allen et al., 2008). 

4. Outdoor air: The CARB data of 84 samples taken in 2004 from seven monitors on 

12 dates from locations in the Bay Area and the South Coast were used for the profile of 

outside air. The profile at 158 pg/m3 might be higher than the profiles measured by Hoh 

and Hites (2005) or Strandberg et al. (2001); in fact, the congener-specific measurements 

made in these two studies in urban areas are similar to the California measurements. For 

example, the CARB data included a measurement of 25 pg/m3 for BDE 209, while Hoh 

and Hites (2005) measured 60.1 pg/m3 as an average of 28 samples in 2002/2003. 

Strandberg et al. (2001) measured 0.30 pg/m3 for BDE 209, but their measurements 

pertain to 1997−1999, which was before the prominent use of the deca formulation. 

Strandberg et al. (2001) did measure concentrations of 33 and 16 pg/m3 for BDE 47 and 

BDE 99, respectively, which compares well with Hoh and Hites (2005), measuring 

17 and 7 pg/m3 for these two congeners, and with the CARB data showing 53 and 

51 pg/m3 for these two congeners, respectively. Therefore, it would appear that the 

CARB data capture current urban conditions, and because 84 measurements of 10 of the 

14 desired congeners were available (more than any other study), these data were used. 
 4-97 Table 4-5. Exposure media concentrations Exposure media Congeners 

17 28 47 66 85 99 100 138 153 154 183 197 206 

Water, pg/L 6.6 3.3 42.7 2.2 1.3 27.6 7.2 0.3 3.9 2.9 4.4 0.1 1.4 4

Surface soil, ng/g dwt --- --- 1.9 <0.1 <0.1 3.6 0.4 --- 5.7 4.8 37.4 12.4 0.8 

House dust, ng/g --- ND 1,857 21 100 2,352 911 181 243 156 60 --- --- 

Outdoor 3 air, pg/m 2 3 53 --- 2 51 13 --- 4 4 1 --- --- 

Indoor air, 3 pg/m 8 27 177 4 3 79 16 --- 5 7 --- --- --- 

Shellfish, ng/g wwt ND ND 3.6 ND ND 1.2 0.9 ND ND ND ND ND ND N

Finfish, ng/g wwt 

0.002 0.01 0.20 0.005 0.0006 0.04 0.04 0.0007 0.006 0.015 0.001 0.0002 0.0004 

Beef, ng/g wwt 

ND 0.000

6 

0.009 ND ND 0.01 0.001 ND 0.003 0.0009 0.0004 ND ND 

Pork, ng/g wwt 

ND 0.000

7 

0.031 ND 0.0006 0.033 0.005 0.0009 0.007 0.003 0.02 0.01 0.0008 

Poultry, ng/g wwt 

ND 0.000

3 

0.005 ND ND 0.007 0.001 ND 0.001 0.001 0.001 ND 0.001 0

Dairy, ng/g wwt 

ND 0.001 0.04 ND 0.0006 0.032 0.005 ND 0.004 0.002 0.001 0.002 0.029 

Eggs, ng/g wwt 

ND ND 0.009 ND ND 0.018 0.007 ND 0.007 0.004 ND ND 0.005 0
4-98

Table 4-5. Exposure media concentrations (continued)

a

For Oros et al. (2005), concentrations reported as “Q” meaning “detected, but not reportable because outside QA limits” not counted in averaging; “bdl” 

meaning below detection limit counted as 0. b

Allen et al. (2006) sampled 3 locations within 20 urban residences, and presented geometric means for the three locations over all 20 homes. Results presented 

are the average of the 3 geometric mean values. c

Oros et al. (2005) claims that only BDEs 47, 99, and 100 were detected in sampling of clams, oysters, and mussels―that others were measured but not 

detected―from SF Estuary.

d

All the samples in Schecter et al (2009) are composites of 10 samples, so 10 times the “n” is represented in the averages presented. In calculation of averages, 

nondetects were set equal to zero due to the range of detection limits presented, including many which exceeded the positive detections. 

--- = no data available; ND = data available but not detected
4-99

5. Indoor air: The only indoor air measurements in the United States that could be 

representative of general population exposures were taken in 20 urban residences in 

Boston, MA (Allen et al., 2007). The geometric mean concentrations were presented for 

three locations (personal, bedroom, living room), and the average of the three geometric 

means were used as the representative indoor air concentrations. The sum of the 

geometric mean concentrations for reported congeners for the three locations equaled 

605 pg/m3 for the “personal” samples, which were taken near the breathing zone in the 

bedroom, 392 pg/m3 in the bedroom, and 366 pg/m3 in the living room. 

6. Shellfish: The only shellfish data available were on clams, oysters, and mussels from the 

San Francisco Estuary (Oros et al., 2005), so these data were used for the profile. While 

measurements were made for all 14 congeners, it was stated that nondetects were found 

for all but BDE 47, BDE 99, and BDE 100. The wwt total average concentration was 

5.7 ng/g wwt, with a range of 2−13 ng/g wwt. 

 

7. Finfish: The retail market place data from Schecter et al. (2009b) were thought to be the 

most representative for use in this exposure assessment. Seven composite (each 

including 10 individual samples of the same fish type) samples were taken, ranging from 

canned tuna to cod to frozen fish sticks. These data showed a small decline from a 

similarly designed retail market place survey from the same research group and 

geographic location. For example, the average concentrations of these four congeners 

(which the highest found in 2006) were: BDE 47 at 0.2 ng/g wwt (it was 0.6 ng/g wwt in 

the earlier survey), BDE 99 at 0.04 ng/g wwt (compared to 0.17 ng/g wwt), BDE 100 at 

0.04 (compared to 0.13 ng/g wwt), and BDE 209 at 0.008 ng/g wt (compared to 0.098 

ng/g wwt). The total concentration (sum of key congeners) declined from 1.2 to 0.3 ng/g 

wwt. There were some substantially higher measurements taken from fish in the Great 

Lakes, including findings by Manchester-Neesvig et al. (2001) on 21 coho and Chinook 

salmon samples showing an average of 80 ng/g wwt, or the temporal study by Zhu and 

Hites (2004) showing an average of 120 ng/g wwt for lake trout in Lakes Superior, 

Michigan, Huron, and Ontario. However, other retail market surveys, such as the one in 

California, show a range of 0.04 to 4.9 ng/g wwt in fish, and the one in Canada, including 

122 fish and shellfish, shows a range of 0.02 ng/g wwt (in shrimp) to 1.6 ng/g wwt (in 

trout). Other data on farmed fish and fish from abroad, reviewed in Section 4.6, similarly 

showed concentrations mostly below 10 ng/g wwt and near the value of 0.3 ng/g wwt 

measured by Schecter et al. (2009b).

8. Beef: The retail market place data from Schecter et al. (2009b) were used for beef 

congener data. The total concentration of 0.028 ng/g wwt, which was the average of 

three composite beef samples (hamburger, roast beef, canned chili), was lower than 

similar retail market basket surveys: Schecter et al (2006a) earlier found an average of 

0.13 ng/g wwt in three samples, Huwe and Larsen’s (2005) beef samples (n = 11) 

averaged 0.42 ng/g lwt (roughly 0.08 ng/g wwt) for the five major congeners (BDE 47, 

BDE 99, BDE 100, BDE 153, and BDE 154), and Luksemburg et al. (2004) beef samples 

(n = 4) from Northern California averaged 0.15 ng/g wwt total. The data of Schecter et al 

(2009b) represents the most recent sampling and his study was comprehensive with 
4-100

regard to congeners measured (24) and food types samples (composites of 10, of 

31 different food types).

9. Pork: The retail market place data from Schecter et al. (2009b) were used for pork 

congener data. The total concentration of 0.14 ng/g wwt, which was the average of 

3 composite samples (of bacon, sausages, and ham) was generally lower than comparable 

studies. For example, the earlier, similarly designed retail market basket survey by 

Schecter et al (2006a) found an average total of 0.28 ng/g wwt in seven pork samples. 

Huwe and Larsen (2005), whose pork samples include 11 bacon and 11 pork fat samples, 

showed a total of 1.59 ng/g lwt for the five major BDE congeners (BDE 47, BDE 99, 

BDE 100, BDE 153, and BDE 154). Assuming about a 15% lipid weight of pork, this 

translates to 0.24 ng/g wwt. The data of Schecter et al. (2009b) represents the most 

recent sampling and his study was comprehensive with regard to congeners measured 

(24) and food types samples (composites of 10, of 31 different food types). 

10. Poultry: The retail market basket data from Schecter et al. (2009b) was used for poultry 

congener data. The total concentration of 2 composite samples (sliced turkey and sliced 

chicken breast) of 0.06 ng/g wwt is generally lower than comparable studies. For 

example, the earlier, similarly designed retail market basket survey by Schecter et al 

(2006a) found an average total of 0.36 ng/g wwt for three poultry samples. Luksemburg 

et al. (2004).found an average of 0.41 ng/g wwt for seven poultry samples (four chicken 

and three turkey samples), and Huwe and Larsen (2005) found an average of 2.78 ng/g 

lwt in 22 chicken fat samples. Assuming 15% fat in whole-weight chicken, this translates 

to about 0.42 ng/g wwt. Interestingly, the duck sample from Luksemburg et al. (2004) at 

2.5 ng/g wwt, and the duck sample from Schecter et al. (2006a) at 1.3 ng/g wwt, were the 

highest of the poultry samples, and these were both not included in the displayed 

averages for their studies. Duck was not included in the recent Schecter et al (2009b) 

study. This high concentration occurred because of the high fat content of duck, listed as 

75% lipid in Schecter et al. (2006a). The lipid content of the chicken and turkey samples 

in the four studies noted were between 2 and 11% lipid, and these wet weight 

concentrations are felt to better represent consumed poultry. The data of Schecter et al 

(2009b) represents the most recent sampling and his study was comprehensive with 

regard to congeners measured (24) and food types samples (composites of 10, of 31 

different food types). 

11. Dairy: The retail market place data from Schecter et al. (2009b) will be used for dairy 

congener data. A total of eight composite samples were taken, including three cheeses, 

whole milk, butter, ice cream, and two yogurts. The highest concentration was found in 

the butter sample, at 6.2 ng/g wwt. However, this sample was 91% fat, so for this 

category, lipid average concentrations were determined from the eight samples, and then, 

exposure was determined based on a whole weight concentration. It was assumed that 

the milk lipid content was 0.018 (1.8%) and that all other dairy had a lipid content of 0.12 

(12%). More detail on the procedure to determine PBDEs via dairy intake is provided in 

Chapter 5. The lipid weight average total PBDE concentration of total PBDEs (sum of 

18 congeners found in sufficient frequency in Schecter et al., 2009b) was 1.3 ng/g lwt, 

which would translate to about 0.16 ng/g wwt assuming 12% lipid content. By 
4-101

comparison, the average over 15 samples in the Schecter et al. (2006) retail market 

survey (with a mean lipid content of 10%) was similar at 0.11 ng/g wwt. The data of 

Schecter et al (2009b) represents the most recent sampling and his study was 

comprehensive with regard to congeners measured (24) and food types samples 

(composites of 10, of 31 different food types). 

12. Eggs: Schecter et al. (2009b) included a composite sample of ten eggs. The total 

concentration of this sample was 0.09 ng/g lwt, which is the same as their reported 

concentration of about 0.09 ng/g wwt in the earlier study (Schecter et al, 2006a). The 

data of Schecter et al (2009b) represent the most recent sampling and his study was 

comprehensive with regard to congeners measured (24) and food types samples 

(composites of 10, of 31 different food types). 

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5-1

5. HUMAN EXPOSURE

5.1. INTRODUCTION

The rise in polybrominated diphenyl ether (PBDEs) in breast milk in the United States

throughout the 1990s and into the 2000s, coupled with the finding that milk concentrations in 

American women exceed those of European women by factors of 10 or more, has served to focus 

attention on U.S. exposures to PBDEs. The first section of this chapter reviews the data on body 

burdens of PBDEs, with a focus on data from the United States. Robust data are available on 

blood and human milk, while limited data are available on other matrices including adipose 

tissue and liver. The section on body burdens concludes with assignment of representative 

PBDE congener background profiles in blood and mother’s milk. The next section reviews 

estimates of dose of PBDEs currently available in the literature. This is followed by 

development of an estimate of background dose using the environmental media concentration 

profiles developed in Chapter 4, in combination with exposure contact rates. The dose estimate 

made here is compared to the literature estimates of dose. The next section of the chapter 

attempts to use a simple pharmacokinetic framework to see if the dose estimates can explain 

body burdens of the individual congeners. Figure 5-1 depicts this approach. In combination 

with mother’s milk concentrations, the same pharmacokinetic (PK) model was used to evaluate 

the impact of breast-feeding on the PBDE body burden of infants. Other exposures of interest, 

including fetal exposures, childhood exposures, and occupational exposures, are examined. The 

chapter concludes with discussions on the uncertainties in the procedures used in this chapter, 

and a series of findings from this examination of U.S. exposures to PBDEs. 

5.2. BODY BURDEN DATA

Data from around the world on PBDE body burdens, with an emphasis on data from the 

United States, are reviewed in this section. Sections on blood and human milk conclude with a 

table displaying individual congener data; data on adipose and other tissues are too sparse to 

warrant a table of raw data. This section on body burden data concludes with a table suggesting 

representative profiles in blood and mother’s milk for selected congeners. 
5-2

. 

Figure 5-1. Approach for characterizing exposure to PBDEs in this report.

5.2.1. Blood Data 

Seventeen studies that contained data on PBDEs in blood in the United States are 

reviewed in this section. Table 5-1 shows congener-specific data from these studies. 

The most statistically rigorous and expansive study of background exposures to PBDEs is 

a recent analysis of 2003/2004 National Health and Nutritional Examination Survey (NHANES) 

data by Sjodin et al. (2008). Unfortunately, brominated diphenyl ether (BDE) 209 was not 

measured in this NHANES study, although Sjodin et al. (2008) note that the concentration of 

BDE 209 from pooled NHANES 2001/2 samples was 2 ng/g lipid weight (lwt; unpublished data; 

no further details supplied). A total of 2,040 serum samples from individuals 12 years of age and 

older were analyzed for 10 BDE congeners, including BDE 17, BDE 28, BDE 47, BDE 66, BDE 

85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183. The geometric mean concentrations 

over the entire population were, in ng/g lwt, in descending order: 20.5 for BDE 47, 5.7 for BDE 

153, 5.0 for BDE 99, 3.9 for BDE 100, and 1.2 for BDE 28. Geometric means were not provided 
5-3

Table 5-1. Blood concentrations of PBDE congeners in Americans

Congener Concentration

(ng/g lwt) Comment Citation

DiBDE

15 0.2; 0.09–0.5 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

TriBDE

17 0.3 N = 1; Pool of 15 individuals from Philadelphia, Memphis, 

Miami

Focant et al. 

(2004)

0.1, 0.03, 0.05 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

0.013 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

0.003 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

0.1, ND–0.7 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

28 1.2 N = 2040; NHANES 2003/2004; geometric mean Sjodin et al. 

(2008)

0.8 N = 1; Pool of 15 individuals from Philadelphia, Memphis, 

Miami

Focant et al. 

(2004)

1.3, 1.9, 1.1 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

0.65 N = 100; median serum level in 100 mothers who were 

pregnant and near WTC on 9/11

Wolff et al. 

(2005)

0.92 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

0.58 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

0.4, 1.0 N = 20 mothers and 20 toddlers pairs; median from 

mother, then child (ages 1.5−4) 

EWG (2008)

0.6; ND Medians from “high” (n = 22) and “low” (n = 14) New 

York State sportfish consumers

Spliethoff et al. 

(2008)

1.3, 0.09–5.6 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

30 0.2; ND–0.06 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

37 <0.1; ND–0.06 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)
5-4

Table 5-1. Blood concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

TetraBDE

47 20.5 N = 2040; NHANES 2003/2004; geometric mean Sjodin et al. 

(2008)

0.6 (<0.4−23.8) Median, range; 6 of 12 blood donor samples collected in 

IL in 1988 quantified

Sjodin et al. 

(2001)

13.2, 0.7−1388.6 Geometric mean; range of 93 urban angler samples in NY 

and NJ; 93% detected

Morland et al. 

(2005)

28.1 N = 1; Pool of 15 individuals from Philadelphia, Memphis, 

Miami

Focant et al. 

(2004)

34 (29−98) Median from 2000/2002 (range) in trend study from 

pooled blood from around United States

Sjodin et al. 

(2004a)

20.8; 0.7–109.0 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

28, 9.1−310 Median; range from maternal/umbilical cord blood 

(n = 12); 2001; umbilical cord blood was identical

Mazdai et al. 

(2003)

50.6 (<10−511) Mean, range from Laotian reproductive age women in San 

Francisco area, 1997−1999

Petreas et al. 

(2003)

32.5, 44.2, 25.0 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

0.6 (<0.4−23.8) Median, range (n = 12) from individual donors collected in 

1988 in Illinois

Sjodin et al. 

(2001)

23−60, 94−137, 

186−245

Range from two parents, one 5 year-old, and one 18-month 

old from case study in CA in 2004

Fischer et al. 

(2006)

9.7 N = 100; median serum level in 100 mothers who were 

pregnant and near WTC on 9/11

Wolff et al. 

(2005)

22.6 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

13.1 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

11, 205 N = 24; median and max from a cohort of pregnant Latina 

women in CA

Bradman et al. 

(2007)

8.8, 30.6 N = 20 mothers and 20 toddlers pairs; median from child, 

then child (ages 1.5−4) 

EWG (2008)

12.8, 1.3 Medians from “high” (n = 22) and “low” (n = 14) New 

York State sportfish consumers

Spliethoff et al. 

(2008)

18, 17 Geometric means for men (n = 350) and women (n = 150) 

from a study of sport fishers

Anderson et al. 

(2008)
5-5

Table 5-1. Blood concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

49 0.1; ND–0.8 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

66 0.3 N = 1; Pool of 15 individuals from Philadelphia, Memphis, 

Miami

Focant et al. 

(2004)

0.3, NA, 0.4 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

ND N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

0.1 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

0.2, ND Medians from “high” (n = 22) and “low” (n = 14) New 

York State sportfish consumers

Spliethoff et al. 

(2008)

0.2; ND–0.9 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

71 <0.1; ND–0.1 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

75 <0.1; ND–0.1 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

77 ND, NA, 0.01 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

PentaBDE

85 1.0, 0.2−109.1 Geometric mean; range of 92 urban angler samples in NY 

and NJ; 27% detected

Morland et al. 

(2005)

0.8 N = 1; Pool of 15 individuals from Philadelphia, Memphis, 

Miami

Focant et al. 

(2004)

0.7 (0.5−1.4) Median from 2000/2002 (range) in trend study from 

pooled blood from around United States

Sjodin et al. 

(2004a

NA, 1.1, 1.2 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

0.4 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

0.3 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

0.3, 5.0 N = 24; median and maximum from a cohort of pregnant 

Latina women in CA

Bradman et al. 

(2007)
5-6

Table 5-1. Blood concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

ND, 0.4 N = 20 mothers and 20 toddlers pairs; median from 

mother, then child (ages 1.5−4) 

EWG (2008)

0.4, ND Medians from “high” (n = 22) and “low” (n = 14) New 

York State sportfish consumers

Spliethoff et al. 

(2008)

0.3; ND–1.8 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

99 5.0 N = 2,040; NHANES 2003/2004; geometric mean Sjodin et al. 

(2008)

0.3, <0.2−3.7 Median, range; 8 of 12 U.S. blood donor samples collected 

in IL in 1988 quantified; 

Sjodin et al. 

(2001)

3.2, 0.3−545.5 Geometric mean; range of 93 urban angler samples in NY 

and NJ; 66% detected

Morland et al. 

(2005)

9.2 N = 1; Pool of 15 individuals from Philadelphia, Memphis, 

Miami

Focant et al. 

(2004)

11 (6.8−26) Median from 2000/2002 (range) in trend study from 

pooled blood from around United States

Sjodin et al. 

(2004a)

5.7, 2.4−68 Median; range from maternal/umbilical cord blood 

(n = 12); 2001; umbilical cord blood was identical

Mazdai et al. 

(2003)

8.4, 12.8, 11.1 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

4−16, 28−34, 

37−45

Range from two parents, one 5 year-old, and one 18-month 

old from case study in CA in 2004

Fischer et al. 

(2006)

1.5 N = 100; median serum level in 100 mothers who were 

pregnant and near WTC on 9/11

Wolff et al. 

(2005)

6.0 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

2.7 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

2.9, 54.0 N = 24; median and maximum from a cohort of pregnant 

Latina women in CA

Bradman et al. 

(2007)

1.5, 6.2 N = 20 mothers and 20 toddlers pairs; median from 

mother, then child (ages 1.5−4) 

EWG (2008)

3.4, ND Medians from “high” (n = 22) and “low” (n = 14) New 

York State sportfish consumers

Spliethoff et al. 

(2008)

5, 5 Geometric means for men (n = 350) and women (n = 150) 

from a study of sport fishers

Anderson et al. 

(2008)

4.3; 0.6–24.2 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)
5-7

Table 5-1. Blood concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

100 3.9 N = 2040; NHANES 2003/2004; geometric mean Sjodin et al. 

(2008)

0.2, <0.1−2.4 Median, range; 10 of 12 U.S. blood donor samples

collected in IL in 1988 quantified

Sjodin et al. 

(2001)

2.7, 0.3−280.6 Geometric mean; range of 93 urban angler samples in NY 

and NJ; 88% detected

Morland et al. 

(2005)

6.9 N = 1; pool of 15 individuals from Philadelphia, Memphis, 

Miami

Focant et al. 

(2004)

5.9 (3.5−18) Median from 2000/2002 (range) in trend study from 

pooled blood from around United States

Sjodin et al. 

(2004a)

4.2, 1.9−110 Median; range from maternal umbilical cord blood 

(n = 12); 2001; umbilical cord blood was identical

Mazdai et al. 

(2003)

5.7, 5.2, 4.7 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 11) from TX; mean of 

39 individuals―28 MS, 10 NY

Schecter et al. 

(2005)

8−22, 30−39, 

57−87

Range from two parents, one 5 year-old, and one 18-month 

old from case study in CA in 2004

Fischer et al. 

(2006)

1.8 N = 100; median serum level in 100 mothers who were 

pregnant and near WTC on 9/11

Wolff et al. 

(2005)

5.7 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

3.7 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

1.8, 44.0 N = 24; median and maximum from a cohort of pregnant 

Latina women in CA

Bradman et al. 

(2007)

1.2, 6.2 N = 20 mothers and 20 toddlers pairs; median from 

mother, then child (ages 1.5−4) 

EWG (2008)

3.0, ND Medians from “high” (n = 22) and “low” (n = 14) New 

York State sportfish consumers

Spliethoff et al. 

(2008)

3.3 Geometric means for men (n = 350) and women (n = 150) 

from a study of sport fishers

Anderson et al. 

(2008)

2.9; 0.1–19.7 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

HexaBDE

138 NA, 0.3, 0.2 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

0.07 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)
5-8

Table 5-1. Blood concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

0.04 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

ND, ND Medians from “high” (n = 22) and “low” (n = 14) New 

York State sportfish consumers

Spliethoff et al. 

(2008))

0.1; ND–0.4 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

153 5.7 N = 2040; NHANES 2003/2004; geometric mean Sjodin et al. 

(2008)

0.3, 0.1−2.0 Median, range; 12 of 12 U.S. blood donor samples 

collected in IL in 1988 quantified

Sjodin et al. 

(2001)

3.2, 0.4−165.2 Geometric mean; range of 93 urban angler samplers in NY 

and NJ; 96% detected

Morland et al. 

(2005)

6.2 Pool of 15 individuals from Philadelphia, Memphis, Miami Focant et al. 

(2004)

7.3 (1.8−17) Median from 2000/2002 (range) in trend study from 

pooled blood from around United States

Sjodin et al. 

(2004a)

2.9, 1−83 Median; range from maternal/umbilical cord blood 

(n = 12); 2001; umbilical cord blood was identical

Mazdai et al. 

(2003)

12.3, 11.7, 5.7 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

19−42, 49−65, 

75−141

Range from two parents, one 5 year-old, and one 18-month 

old from case study in CA in 2004

Fischer et al. 

(2006)

1.8 N = 100; median serum level in 100 mothers who were 

pregnant and near WTC on 9/11

Wolff et al. 

(2005)

14.6 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

8.9 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

1.5, 35 N = 24; median and maximum from a cohort of 

24 pregnant Latina women in CA

Bradman et al. 

(2007)

5.8, 15.3 N = 20 mothers and 20 toddlers pairs; median from 

mother, then child (ages 1.5−4) 

EWG (2008)

3.0, 0.7 Medians from “high” (n = 22) and “low” (n = 14) New 

York State sportfish consumers

Spliethoff et al. 

(2008)

6.6 Geometric means for men (n = 350) and women (n = 150) 

from a study of sport fishers

Anderson et al. 

(2008)

5.2; 0.1–20.9 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)
5-9

Table 5-1. Blood concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

154 0.6, 0.09−24.7 Geometric mean; range of 89 urban angler samples in NY 

and NJ; 25% detected

Morland et al. 

(2005)

1.2 N = 1; pool of 15 individuals from Philadelphia, Memphis, 

Miami

Focant et al. 

(2004)

0.95 (0.5−1.8) Median from 2000/2002 (range) in trend study from 

pooled blood from around United States

Sjodin et al. 

(2004a)

0.3 (ND−6.1) Median, range from maternal/umbilical cord blood 

(n = 12); 2001; umbilical cord blood was identical

Mazdai et al. 

(2003)

0.8, 0.8, 1.0 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY

Schecter et al. 

(2005)

0.6−6, 3−7, 4−17 Range from two parents, one 5 year-old, and one 18-month 

old from case study in CA in 2004

Fischer et al. 

(2006)

0.6 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

0.4 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

0.3, 4.2 N = 24; median and maximum from a cohort of 

24 pregnant Latina women in CA

Bradman et al. 

(2007)

0.6, ND Medians from “high” (n = 22) and “low” (n = 14) New 

York State sportfish consumers

Spliethoff et al. 

(2008)

0.3; 0.05–1.5 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

155 0.1; ND–0.2 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States. 

PSR (2009)

HeptaBDE

183 0.2, 0.1−1.3 Median, range; 12 of 12 U.S. blood donor samples 

collected in IL in 1988 quantified

Sjodin et al. 

(2001)

0.5, 0.1−2.0 Geometric mean; range of 93 urban angler samplers in NY 

and NJ; 29% detected

Morland et al. 

(2005)

ND, 0−2.7 Median; range from maternal/umbilical cord blood 

(n = 12); 2001; umbilical cord blood was identical

Mazdai et al. 

(2003)

0.3, 0.4, 0.4 Three 2003 results: one pooled serum (n = 100) from TX, 

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

0.04 N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

0.05 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)
5-10

Table 5-1. Blood concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

ND, ND N = 20 mothers and 20 toddlers pairs; median from 

mother, then child (ages 1.5−4) 

EWG (2008)

0.3; 0.05–0.8 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States.

PSR (2009)

190 <0.1; ND–0.2 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

OctaBDE

197 0.3, 0.5 N = 20 mothers and 20 toddlers pairs; median from 

mother, then child (ages 1.5−4) 

EWG (2008)

203 ND(0.1), <0.1−0.2 Median, range; 5 of 12 U.S. blood donor samples collected 

in IL in 1988 quantified

Sjodin et al. 

(2001)

0.2; 0.07–0.3 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States.

PSR (2009)

NonaBDE

206 0.4; 0.2–0.8 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

207 0.6; 0.4–1.0 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

208 0.3; 0.2–0.7 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

DecaBDE

209 ND(1.0), 

<1.0−33.6

Median, range; 5 of 12 U.S. blood donor samples collected 

in IL in 1988 quantified

Sjodin et al. 

(2001)

NA, 1.4, 2.7 Three 2003 results: one pooled serum (n = 100) from TX,

one pooled whole blood (n = 100) from TX; mean of 

39 individuals―29 MS, 10 NY 

Schecter et al. 

(2005)

2−23, 9−143, 

19−233

Range from two parents, one 5 year-old, and one 18-month 

old from case study in CA in 2004

Fischer et al. 

(2006)

ND N = 8; mean from vegans (ND = 0) Schecter et al. 

(2006a)

3.7 N = 11; mean from blood/milk study (ND = 0) Schecter et al. 

(2006c)

ND, 1.7 N = 20 mothers and 20 toddlers pairs; median from 

mother, then child (ages 1.5−4) 

EWG (2008)

3.1; ND–9.0 Mean (ND = 0), range; samples from 20 health care 

providers from 10 states around the United States

PSR (2009)

ND = not detected. 
5-11

for the other BDE congeners because they were detected at less than 60%, including the low 

frequencies of 5%, 15%, 21%, and 23% for BDE 17, BDE 183, BDE 66, and BDE 85, 

respectively. The sum of these geometric means was 36.3 ng/g lwt. The 95th percentile for the 

sum of PBDEs was 291 ng/g lwt, and the maximum found was 3,680 ng/g lwt, with BDE 47 at 

2,350 ng/g lwt for this individual. A statistically significant relationship between age and 

concentration was found for BDE 28, BDE 47, BDE 99, BDE 100, and BDE 153. Specifically, 

the highest concentrations were found in the age group 12−19 years, with lower concentrations 

for 20−39, and 40−59, but then a rise for the category >60 years. For example, the geometric 

mean concentrations of BDE 47 for these four age categories, respectively, were 28.2, 21.5, and 

17.7, and then rose to 19.7 ng/g lwt. Although not statistically significant, the geometric mean 

concentrations for males were higher than females for BDE 47, BDE 99, BDE 100, BDE 153, 

and BDE 154. Although BDE 47 was found most frequently and at the highest concentrations 

for the survey as a whole, 10.5% of participants had BDE 153 concentrations that were higher 

than BDE 47. Race was identified as a statistically significant factor for BDE 99; the geometric 

mean for non-Hispanic blacks was the highest at 6.2 ng/g lwt, Mexican Americans were at 5.9 

ng/g lwt, and non Hispanic whites were at 4.7 ng/g lwt. 

Other than NHANES, the most comprehensive study of PBDEs in blood was conducted 

by Schecter et al. (2005). Although limited geographically, this study contained two pooled 

samples of 100 individuals each and 39 samples of blood from individuals in Mississippi and 

New York. In addition, the study included an archived blood sample from 1973, which was 

from a pool of 100 individuals from Dallas, Texas. All samples were analyzed for 13 congeners 

including BDE 17, BDE 28, BDE 47, BDE 66, BDE 77, BDE 85, BDE 99, BDE 100, BDE 138, 

BDE 153, BDE 154, BDE 183, and BDE 209. All congeners were nondetects for the 1973 

sample. The two pooled samples collected in 2003, one serum and one whole blood, both 

n = 100 as noted, were from discarded samples from the University of Texas Southwestern 

Medical Center in Dallas. The 39 individuals sampled in 2003 included 29 from Mississippi and 

10 from New York. The results from the three groups of samples from 2003 were similar: total 

concentrations of the 13 congeners were 61.8 ng/g lwt for the serum pool, 79.7 ng/g lwt for the 

whole blood pool, and 52.6 ng/g lwt as the mean for the 39 individuals. The congener-specific 

trends were similar as well: BDE 47 dominated the profile by encompassing between 44 and 

53% of the total, with BDE 99 and BDE 153 comprising about the same amounts―BDE 99 was 
5-12

between 14 and 21% and BDE 153 was between 11 and 20%. From the individual samples, an 

interesting trend was that women had higher concentrations than men did: the range and mean of 

total BDE in 22 men were 4.6−192.8 ng/g lwt and 25.1 ng/g lwt., and for 17 women, the range 

and mean of total BDE were 5.6−365.5 ng/g lwt and 74.1 ng/g lwt. Although women had a 

higher level of PBDEs in their blood than men in this study, Schecter et al. (2005) stated that 

their results were not statistically significant. BDE 209 was found at low levels in the pooled 

blood, 1.4 ng/g lwt, and in the individual samples, it averaged 1.7 ng/g lwt with nondetects in 

19 of 39 individual samples. 

The public interest group, Physicians for Social Responsibility, measured a suite of 

contaminants in the blood of 20 volunteers in the medical sector (nurses, physicians, and others) 

from 10 states around the country (PSR, 2009). They measured for perfluorinated compounds 

and PBDEs in blood, and phthalate metabolites, triclosan, bisphenol A, and mercury in urine. 

They sampled between February and April of 2009, and measured 28 BDE congeners from 

BDE 15 to BDE 209. Assuming ND = 0, they found an average total PBDE concentration of 

41.2 ng/g lwt, dominated by BDE 47 at 20.8 ng/g lwt. BDE 209 was found in half the samples, 

with an average concentration of 3.1 ng/g lwt (assuming again ND = 0). 

 Sjodin et al. (2004a) conducted a more rigorous temporal evaluation by collecting 

samples that represented different time frames from the mid-1980s until the early 2000s. 

Specifically, serum pools were collected in the southeastern United States from a blood bank in 

Memphis, TN, representing years 1985−1997, and 2002, and serum pools were collected in 

Seattle, WA, representing years 1999−2002. The pooled samples were clustered to represent the 

following time frames: 1985−1989 (n = 9), 1990−1994 (n = 14), 1995−1999 (n = 10), and 

2000−2002 (n = 7), and the samples were measured for BDE 47, BDE 85, BDE 99, BDE 100, 

BDE 153, and BDE 154. A clear trend was seen, with total concentrations rising from 

9.6 ng/g lwt in the 1985−1989 time frame, to 48 ng/g lwt in 1990−1994, 71 ng/g lwt in 

1995−1999, and 61 ng/g lwt in 2000−2002. BDE 47 comprised 55−65% of the total in the four 

time frames. In contrast to BDEs, BB-153 (2,2’,4,4’,5,5’-hexabromobiphenylhexabrominated 

biphenyl, a marker for polybrominated biphenyls) and CB-153 (2,2',4,4',5,5'-

hexachlorobiphenyl, a marker for polychlorinated biphenyls) decreased in the samples over time. 

 Sjodin et al. (2001) provided another data set from the latter part of the 1980s showing 

similarly low concentrations of PBDEs. Twelve samples from U.S. donors who provided blood 
5-13

at a commercial blood collection facility in the state of Illinois in 1988, and then stored at −70°C, 

were retrieved for analysis. Seven congeners were quantified: BDE 47, BDE 99, BDE 100, 

BDE 153, BDE 183, BDE 203, and BDE 209. Three unidentified octa BDE and three 

unidentified nona BDEs were noted. The data were reported in units of pmol/g lwt, converted to 

ng/g lwt by multiplying by the congener’s molecular weight in pg/pmol, and then pg converted 

to ng by a multiplication of 0.001 ng/pg. Although total concentrations were not provided―only 

median and range of concentrations for individual congeners were provided―an estimate of a 

median total concentration developed as a sum of the medians of individual congeners was 

2.7 ng/g lwt. This is much lower than the 50−80 ng/g lwt total found by Schecter et al. (2005) 

from samples taken in 2003. Of interest is the finding of BDE 209 in 5 of 12 samples, with 

positives ranging from 1.5−33.6 ng/g lwt. Sjodin et al. (2001) stated that BDE 209 has a short 

half-life in humans, 6.8 days, so the presence here suggests continual exposure near the time of 

blood sample collection. Although not stated by the authors, its presence could also be 

laboratory contamination. Also of note is that the congener most consistently found of the other 

6 was BDE 153, found at a range of 0.1−2.0 in all 12 samples. 

 Adipose tissue and serum were sampled from two disparate cohorts of women who were 

sampled in the late 1990s, and they were compared with a third cohort of women sampled 

between 1959 and 1967 (Petreas et al., 2003). One set of 32 adipose tissue samples were from 

women undergoing surgery for breast cancer between 1996 and 1998. The second set was serum 

from a group of 50 Laotian women of reproductive age living in the San Francisco Bay area, 

taken in the 1997−1999 time frame. The final set was serum from a study of pregnant women 

enrolled in a case-control study of cryptorchidism and hypospadias as part of the Child Health 

and Development Studies (CHDS), taken between 1959 and 1967. Only BDE 47 could be 

quantified in this study, so it was the only congener measured and reported. Like the Schecter et 

al. (2005) finding of mostly nondetects in blood sampled from 1973, the entire set of 420 serum 

samples from 1959−1967 did not have any detections of BDE 47 at its high quantitation limit of 

10 ng/g lwt. Since PBDEs were not produced commercially until the 1970s, this is not an 

unexpected result. The mean and median from the adipose samples were 29.9 and 16.5 ng/g lwt, 

respectively, with 100% quantifiable measurements. The reproductive study revealed a mean 

and median of 50.6 and 10 ng/g lwt with 48% quantified. There was no relationship between 

BDE 47 concentrations and age in both the adipose tissue and reproductive studies. This 
5-14

contrasts with their finding of an increase of PCB 153 (which was also measured in these 

samples) with age in the adipose and reproductive studies. 

Ninety-three anglers who had sufficient blood volume and who completed a fish 

consumption questionnaire were sampled between 2001 and 2003 (Morland et al., 2005). The 

urban anglers were from New York and New Jersey. Analysis was conducted for BDE 47, 

BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183. BDE 209 could not be reported 

because of high background contamination during the processing of the unknown samples. Fish 

eaters were categorized as “none” (eating no locally caught fish) or “any” (eating some locally 

caught fish), and “any” was further subcategorized based on amount of fish meals per month of 

locally caught fish. The highest congener found was BDE 47, at a geometric mean of 

13.3 ng/g lwt, followed by BDE 99, at 3.2 ng/g lwt, BDE 153 similarly at 3.2 ng/g lwt, and 

BDE 100 at 2.7 ng/g lwt. All other congeners were not detected (ND) or very infrequently 

detected with geometric means less than 1 ng/g lwt. Although total concentrations found were 

not discussed, a sum of the geometric means of the congeners was 24.5 ng/g lwt. The arithmetic 

average might be higher because of the presence of a few very high concentrations. BDE 47 was 

found at a high of 1,388 ng/g lwt, and the high BDE 99 concentration was 546 ng/g lwt, for 

example. There were moderate, but statistically insignificant, increases in BDE concentrations 

from no local fish intake to >1 meal/week. 

 In contrast, Spliethoff et al. (2008) found a significant correlation between levels of BDE 

congeners in the blood of New York state anglers and the frequency with which they consumed 

fish from Lake Ontario and it tributaries. Samples were obtained from 38 individuals who were 

part of a larger New York State Angler Cohort Study whose blood was sampled between 1995 

and 1997. Statistically significant relationships were noted between blood concentrations of 

BDE congeners (BDE 28, BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, and BDE 154) and 

total BDE (sum of congeners) and number of years consuming sportfish between 1980 and 1990, 

and number of meals eaten in the previous year to the analysis. When separating “high” 

consumers (median of 7−11 sportfish meals, maximum of 24−36 meals, in the previous year) 

from “low” consumers (median of 0 meals, maximum of 1−6 meals), the median total 

concentration of high consumers (n = 22) was 23.1 ng/g lwt, while the median for low consumers

(n = 14) was 1.9 ng/g lwt. BDE 47 was the highest congener found, at a median of 12.8 ng/g lwt 

for “high” consumers and 1.9 ng/g lwt for “low” consumers. 
5-15

 In a third study on anglers consuming fish from the Great Lakes, Anderson et al. (2008) 

also found statistically significant relationships between total PBDE levels in blood (sum of 

BDE 28, BDE 47, BDE 49, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153; BDE 209 

measured but not detected) and years of sportfish consumption as well as catfish and shellfish 

consumption. Multivariate as well as highest quartile associations were also found with 

variables such as computer use as well as sportfish consumption. This study entailed collection 

of blood from 508 volunteers (350 men, 158 women) from a full cohort of 4,107 frequent and 

infrequent consumers of Great Lakes fish who were sampled in 2004 and 2005. The geometric 

mean total PBDE from this sample was 27 ng/g lwt for men and 25 ng/g lwt for women, of 

which BDE 47 comprised almost two-thirds. Anderson et al. (2008) compared these results to 

derived geometric means for Caucasian men and women over 30 years from NHANES 

2003/2004 and found the NHANES results to be higher: 34 ng/g lwt for men and 30 ng/g lwt for 

women. 

 Focant et al. (2004) describe a unique methodology for analysis of PBDEs and other 

PBTs. This method uses comprehensive two-dimensional gas chromatography and isotope 

dilution time-of-flight mass spectrometry (GC/GC-IDTOFMS) for the simultaneous 

measurement of selected polychlorinated biphenyls (PCBs), organochlorine pesticides, and 

brominated flame retardants. Unlike classic GC/MS, this method evaluates all contaminants 

simultaneously with one injection into the GC column. Three milk samples and one blood 

sample were analyzed by Focant et al. (2004). The blood sample was from a pooled sample 

collected from 15 individuals in 2002 in three cities: Philadelphia, Memphis, and Miami. 

BDE 17, BDE 28, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 153, and BDE 154 were 

analyzed. The blood sample results were expressed in pg/g wet weight (wwt) from blood serum 

and were only presented on a graph. The results were estimated from the graph and converted to 

ng/g lwt, assuming 0.65% lipid in the serum. The total BDE from one sample was 52.7 ng/g lwt, 

dominated by BDE 47 at 28.1 ng/g lwt, with comparable contributions by BDE 99, BDE 100, 

and BDE 153 at 9.2, 6.9, and 6.2 ng/g lwt, respectively. 

Wolff et al. (2005) conducted a study of exposures among pregnant women who were 

near the World Trade Center (WTC) Site on 9/11/2001. The study involved a complex 

evaluation of exposures, including measurement of key persistent contaminants including 

polycyclic aromatic hydrocarbons, PCBs, dioxins, and PBDEs in blood. The authors did not find 
5-16

an association of PBDEs and measures of potential exposure to WTC contaminants, and they 

generally found low levels of PBDEs. Of 100 mothers, they found median levels for BDE 28, 

BDE 47, BDE 99, BDE 100, and BDE 153 at 0.65, 9.7, 1.5, 1.8, and 1.8 ng/g lwt, respectively.

 Bradman et al. (2007) reports on a sampling of blood from pregnant Latina women, 

primarily from Mexico, living in Salinas Valley, California. The serum specimens were 

collected between September 1999 and January 2001. Seven congeners were measured, 

including BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183. The 

resulting concentrations were low, although not as low as the Wolff et al. (2005) study on 

pregnant women during the World Trade Center attacks. The median total from the 24 women 

was 21 ng/g lwt―with a high of 320 and a low of 5.3 ng/g lwt. Like other studies, the dominant 

congener was BDE 47. The authors showed that the levels are highest among women who have 

spent less than 5 years in the United States, compared to women who have spent more than 

5 years in the United States.

Fischer et al. (2006) present a case study of PBDEs in a family that showed somewhat 

elevated levels in the parents but higher levels in one child and still higher levels in the toddler of 

the family. Samples were collected from a family of four, including 35 and 37 year-old parents, 

a 5-year old daughter, and an 18-month old son in September and December of 2004. The sum 

of BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 in the parents ranged between 64 and 

147 ng/g lwt in the two sampling dates, and BDE 209 contributed a relatively small addition at 

between 2 and 23 ng/g lwt. The story was much different with the children. The 5-year-old 

daughter had concentrations of 237, 239/249 ng/g lwt (the last two were duplicates of the same 

December sample) of the five congeners for the September and then December samples but a 

disparate range of 143 ng/g lwt of BDE 209 in the September sample and 9/12 ng/g lwt in the 

December sample (duplicates). The toddler had the highest concentrations of all: 418 and 

488/476 ng/g lwt for the five congeners and 233 and 19/26 ng/g lwt of BDE 209 in the 

September and December samples, respectively. The authors discounted laboratory error and 

attributed the higher concentrations in the children to exposure to house dust. While the authors 

have discounted laboratory error, it would appear that a decline by an order of magnitude in both 

the toddler and infant is substantial, and could be due to some difference in the two laboratories. 

A decline of this magnitude is plausible because of the short half-life of BDE 209 in humans―it 

has been quantified at values below 10 days, while the half-lives of lower-brominated congeners 
5-17

has been quantified on the order of years. Whether due to differences in laboratories or 

differences in exposures, BDE 209 was quantified by both laboratories, and this in itself is 

worthy of reporting. The higher levels of the other congeners in the toddler were attributed to his 

consumption of human milk, although Fisher et al. (2006) suggest that it might also be due to 

exposure to house dust. Of the non-BDE 209 congeners, the typical trend of seeing BDE 47 at 

the highest concentrations was true for all participants and sampling dates; BDE 153 was second 

most prevalent, ranging from one-third to one-half of the concentration of BDE 47. 

 A second and more recent study providing data on PBDE concentrations in young 

children was conducted by the Environmental Working Group (EWG, 2008). Twenty pairs of 

mothers and children from 16 states were sampled for 20 congeners; positive occurrences were 

found for these 10: BDE 28, BDE 47, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, 

BDE 183, BDE 197, and BDE 209 (no detections for BDE 17, BDE 66, BDE 71, BDE 196, 

BDE 201, BDE 203, BDE 206, BDE 207, and BDE 208; BDE 154 was not included because it 

coeluted with PCB 153). The children ranged in age from 1.5 to 4 years old. The median total 

concentration (sum of the 10 congeners found) from the 20 children was nearly four times that of 

the mother, at 69.3 ng/g lwt for the children versus 18.2 ng/g lwt for the mother. BDE 209 was 

found in 13 of 20 children, with a median of 1.7 ng/g lwt, while it was found in only 9 of 20 

mothers (median at ND). The congener found at the highest concentration was BDE 47 

(medians of 30.6 ng/g lwt in children and 8.8 ng/g lwt in mothers), followed by BDE 153 

(medians of 12.5 ng/g lwt in children and 5.8 in mothers). 

Schecter et al. (2006a) measured the concentrations of 12 BDEs (BDE 17, BDE 28, 

BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, and 

BDE 209) in the blood of eight vegans (no animal food products including dairy). They found a 

range of BDE concentrations of 12.4 to 127 ng/g lwt total, with a median of 23.9 and a mean of 

53.3 ng/g lwt. BDE 47 was the most abundant congener, with the highest mean concentration of 

23 ng/g lwt. The second highest was BDE 153 at 14 ng/g lwt. BDE 99 and BDE 100 totaled 

about 6 ng/g lwt. BDE 209 was not detected, although detection limits (DLs) were relatively 

high at between 2 and 7 ng/g lwt. The authors characterize these findings as lower, but not 

substantially lower, than other studies of BDEs in the blood of Americans, and they suggest that 

this could be the result of not consuming food of animal origin. They note that for dioxins and 

other POPs, foods of animal origin have been attributed as the main source of exposure, and 
5-18

because the concentrations of BDEs in vegans, who have been so for a minimum of 5 years, is 

not that much lower than other populations, exposures other than food may be important for this 

class of compounds. 

Schecter et al. (2006c) provided important insight as to the relative partitioning of PBDEs 

between blood and milk. Eleven mothers donated samples of their blood and milk, which were 

simultaneously measured for BDE 17, BDE 28, BDE 47, BDE 66, BDE 85, BDE 99, BDE 100, 

BDE 138, BDE 153, BDE 154, BDE 183, and BDE 209. The location was not specified, but it 

was assumed to be in Texas, where the primary author is located (and within the article, it was 

stated that the specimens were packed in dry ice and shipped to the analytical lab in 

Germany―coauthors on the paper originate from Germany), and the date of sampling is not 

specified, but is assumed to be 2004 or 2005, since the paper was published in 2006. The focus 

of the paper was the relationship between concentrations found in blood and milk. Tables 5-1 

and 5-2 provide average concentrations of the congeners in blood and milk, and the sum of these 

congener averages (the article did not provide total concentrations or statistics for totals or 

individual congeners; congener-specific averages were generated for this report) is about the 

same in milk and blood at about 33 ng/g lwt. However, there were differences in partitioning 

between blood and milk among the congeners, most importantly for BDE 209 as compared to 

other congeners. The average concentration in blood was 3.7 ng/g lwt, while in milk it was 0.42. 

For other congeners, the concentrations were more nearly similar. The congener found at the 

highest average concentration in blood (B) and milk (M) was BDE 47 at 13.1 ng/g lwt (B) and 

17.1 ng/g lwt (M); the second highest was BDE 153 at 8.4 ng/g lwt (B) and 7.0 ng/g lwt (M). 

Schecter et al. (2006c) generated ratios of blood: milk concentrations; when nondetected, the 

value was set at 1/2 DL. The highest ratio (greatest disparity between blood and milk, with 

blood much higher) was with BDE 209, with the median and mean over the 11 individuals in the 

35−40 range. The ratio was second highest for BDE 183, with median and mean ratios of 1.8 

and 2.3. All other median and mean ratios for the congeners frequently detected (BDE 47, 

BDE 99, BDE 100, BDE 153, and BDE 154) were between 0.7 and 1.5. These data support the 

argument that BDE 209 does not partition equally to all body lipids. 

 A second study providing some insight into partitioning of BDE congeners between 

blood and milk is LaKind et al. (2009). In this study, a group of 10 women were sampled post 

partum, at several intervals and for both blood and milk. Milk was sampled 4 times (1, 2, 
5-19

3 months, and at cessation of breast-feeding) and blood sampled twice (1 month and final milk 

sampling event). Contaminants measured included dioxins and furans, PCB, and PBDEs. 

Summary statistics were provided in LaKind et al. (2009) for BDE 47, BDE 153, and BDE 100. 

The median concentrations over all milk samples (n = 35, 100% detection) for the three 

congeners was 26, 4, and 3.5 ng/g lwt for BDE 47, BDE 100, and BDE 153, respectively. The 

median concentration for the blood samples (n = 17, 100% detection for BDEs 47 and 100; 59% 

for 153) was 15.5, 2.4, and 4 ng/g lwt for BDE 47, BDE 100, and BDE 153, respectively. In 

studying the relationship between blood and milk concentrations, only concurrently measured 

samples collected one month post partum were studied, and only BDE 47 and BDE 100 were 

included because they were detected in all samples. Individual results were not available, but in 

a subgroup of all contaminants measured which includes these BDE congeners (and also 

pesticides/metabolites and 9 PCB congeners), a best-fit line was found at, milk = 1.74 * serum 

(r2 = 0.98), where milk and serum were both expressed on a lipid basis. This suggests that for 

these two BDE congeners, milk concentrations tended to be just under twice the concentration in 

blood. 

 These data suggest a range of approximately 30−100 ng/g lwt total BDEs is

representative of blood in the general population of Americans in the 2000s, with occasional 

measurements in the hundreds and thousands of ng/g lwt. 

 In contrast to the North American studies above, most studies from Europe, Asia, and 

elsewhere suggest concentrations of total BDEs in blood less than 10 ng/g lwt. Studies below 

include the countries of Australia, Sweden, Norway, New Zealand, the United Kingdom (UK), 

the Faroe Islands, Japan, and Nicaragua.

Toms et al. (2008) report on results from pooled samples collected from 8,132 residents 

of Australia in 2002/2003 and 2004/2005. The pools were constructed to correspond to ages 

0−4, 5−15, 16−30, 31−45, 46−60, and >60 years old. A total of 35 congeners were measured, 

making this study the most comprehensive found in terms of the number of congeners measured, 

although BDE 209 was not included. Looking only at results for all age categories greater than 

16 years old (adults essentially), the mean total PBDE concentration for 2002/2003 was 15 

(standard deviation +5, median=13) ng/g lwt, and for 2004/2005, it was 18 (S.D.+5, median=16) 

ng/g lwt. Like the Fischer et al. (2006) study, there was a clear age effect in the data, with the 

highest concentrations found in the youngest individuals. For 2004/2005 (when all age group 
5-20

data were available), the mean was 73 (S.D.+7, median=75) ng/g lipid for 0−4 years of age, 28 

(S.D.+7, median=29) for 5−15 years of age, 20 (S.D.+5, median=18) ng/g lipid for 16−30 years 

of age, and 15−18 ng/g lipid for age categories above 30. 

 Blood from 37 Swedish men were sampled in 1991 and 2001 and measured for BDE 28, 

BDE 47, BDE 99, BDE 100, BDE 128, BDE 154, BDE 153, BDE 183, BDE 196, BDE 197, 

BDE 203, BDE 206, BDE 207, BDE 208, and BDE 209 (Jakobsson et al., 2005). An additional 

10 men were sampled in 1988 and 2002. These men were specifically selected to represent 

levels of fish consumption, so they do not necessarily represent a cross-section of the average 

population. The median and range of total concentrations from 1991 and 2001 were 11 (3.3−59) 

and 14 (4.2−57), respectively, expressed in units of pmol/g lwt. Because individual results were 

not provided, these totals could not be converted to ng/g lwt. However, the conversion would 

result in concentrations that are somewhere between about one-fourth and three-fourths the listed 

concentrations, as the conversion factors for individual congeners range from 0.486 ng/g/pmol/g 

for BDE 47 to 0.959 ng/g/pmol/g for BDE 209. BDE 47 was the dominant congener in 1991, 

which explains about 19% of total concentrations. BDE 154 and BDE 209 explained 12%, and 

BDE 153 accounted for 8%. BDE 153, however, was the dominant congener in 2001: 21% of 

the mean. BDE 47 explained 11%, and BDE 209 and BDE 154 accounted for 10% and 8%, 

respectively. The finding of an increase in BDE 153 and the meaningful contribution of 

BDE 209 are noteworthy in this study. The authors note a decline of other POPs like PCB-153, 

p,p’-DDE, and hexachlorobenzene between 1991 and 2001 of between 30−50%, while no 

decrease and a small increase was noted for BDEs. A second study in Sweden (Karlsson et al., 

2007) looked at levels in air, dust, and blood from individuals in five households. 

Concentrations of the tri-hexa brominated congeners (BDE 28, BDE 58, BDE 66, BDE 99, 

BDE 100, BDE 153, and BDE 154) were near the detection limit for four of the five individuals, 

with concentrations below 10 ng/g lwt for all measurements and the tri-hexa total at less than 

15 ng/g lwt for all individuals. BDE 209 was detected in four of five individuals, at 

concentrations ranging from about 9.4 to 17.4 ng/g lwt. One individual also had quantified 

measurements of BDE 197/204 (coeluting), BDE 196, BDE 206, and BDE 207 at concentrations 

ranging from 3.5 to 9.7 ng/g lwt. While the levels overall are generally low compared to levels 

found in the United States, they are noteworthy like the other study from Sweden discussed here, 

in that the profiles had significant contributions from BDE 209 and even other nona- and 
5-21

decaBDE congeners. A positive relationship was found when plotting a sum of BDE 

concentrations (including BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154, but not 

BDE 209) for each individual against the sum BDE dust concentration found in that individual’s 

home; however, that relationship was driven by one individual having the highest plasma and 

dust sum BDE concentrations. This relationship did not hold up when adding BDE 209 to the 

dust and plasma sumBDE. There was no similar relationship found between air and plasma 

concentrations. 

 Pooled samples of about 20 individuals each from five hospitals in Norway (total number 

of samples analyzed was not provided) were analyzed for 11 BDEs: BDE 28, BDE 37, BDE 47, 

BDE 85, BDE 99, BDE 100, BDE 119, BDE 138, BDE 153, BDE 154, and BDE 183 (Thomsen 

et al., 2005a). Samples for the years 1977, 1982, 1988, 1991, 1994, 1997, 1998, 1999, 2000, 

2001, 2002, and 2003 were obtained. The sum of the seven most abundant congeners (BDE 28, 

BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183) showed a concentration range 

of 0.5 to 5.5 ng/g lwt, with a clear trend of low concentrations for the early years: 0.5 in 1977, 

1.3 in 1982, etc., which consistently rose to high levels between 3.6 to about 5.5 ng/g lwt 

between 1997 and 2003. A spike of about 5 ng/g lwt in 1991 could not be explained. BDE 47 

was found in the highest concentration, except for one of the samples; the relative amount of 

BDE 153 appears to be increasing over time. Results for BDE 209 were presented, but the 

article is not consistent in its reporting of BDE 209 concentrations. The text suggests a median 

BDE 209 of 8.7 ng/g lwt, and the figures in the article show a remarkably high measurement of 

35 ng/g lwt in 2000. These levels appear so much higher than others that even the authors 

suggest, “contamination of the sample cannot be totally excluded.” Also reported in this study 

are age trends for pooled samples stratified by age for age ranges 0−4, 5−14, 15−24, 25−59, and 

>60. Data were presented for 1998 and 2002, with the 1998 data from an earlier study by the 

same team of researchers (Thomsen et al., 2002). The youngest age group, 0−4 years, had the 

highest concentration, at total concentrations at about 11.5 ng/g lwt in 1998 and 10 ng/g lwt in 

2002, while other results for both years and age ranges were between about 3 to 6 ng/g lwt. 

 Harrad and Porter (2007) report on concentrations of congeners including BDE 47, 

BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183 in the blood of 23 individuals (10 males, 

13 females ages 20 to 64) sampled in 2001 in Wellington, New Zealand. The mean 

concentration of total BDEs was similar to other European studies at 7.17 ng/g lwt. Also similar 
5-22

to other studies was the finding that BDE 47 concentrations dominated, accounting for over 50% 

of the concentration. The second most found congener was BDE 153, explaining between 

15 and 20% of concentrations. The authors claim the exposure was due to imported consumer 

goods, because these products are not produced in New Zealand. 

Thomas et al. (2006) measured BDEs in the blood of 154 volunteers in 13 locations in the 

United Kingdom in 2003. They measured 22 congeners including BDE 209. The median total 

BDE was 5.6 ng/g lwt, with a range of 0.6 to 420 ng/g lwt, and only 5% greater than 30 ng/g lwt. 

BDE 209 was quantified in only 11 samples, with a high of 240 ng/g lwt, although the detection 

limit was high at 15 ng/g lwt. BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183 

were regularly detected, similar to other studies, but the median concentration of BDE 153 was 

the highest at 1.7 ng/g lwt, followed by BDE 47 at 0.82 ng/g lwt. This is atypical because

BDE 47 is most often the highest found in profiles. 

Blood concentrations of pregnant Faroese (the Faroe Islands are between Shetland and 

Iceland) women were determined from samples taken in 1994, and then their children’s blood 

was sampled and measured 7 years later, in 2002 (Fangstrom et al., 2005). Fifty-seven mothers 

and 42 children were sampled, of which 41 were mother/child pairs. BDE 47, BDE 99, 

BDE 100, BDE 153, BDE 209, and BDE 153/154 were measured. Concentrations were low, 

with a median total concentration of just over 5 ng/g lwt for both mothers and children. The 

predominant congener for the mother was BDE 47 and the coeluting BDEs 153/154, accounting 

for 26% each. For children, the predominant congener was BDE 153, explaining about 46% of 

the total concentration. BDE 209 was present in both mothers and children at low concentrations 

of 0.8 and 1.0 ng/g lwt, respectively. 

 The study that presented blood data outside of the United States with concentrations 

comparable to those found in the United States was a study in Nicaragua (Athanasiadou et al., 

2008; results are expressed in pmol/g lwt; results expressed in ng/g lwt are found in an earlier 

publication: Faldt et al., 2005). Five pools of serum from teenagers who lived and/or worked 

near a waste disposal in Managua, Nicaragua, and then four pools of serum from women in 

different settings (urban areas, fishing villages, etc.) were analyzed (one analysis per pool). 

BDE 47, BDE 99, BDE 100, BDE 153, BDE 183, BDE 203, and BDE 209 were analyzed. The 

pool of teenagers who both worked at the disposal site and lived nearby (no other pool was that 

exposed) had the highest concentrations, with a total over 600 ng/g lwt. The average of the other 
5-23

eight pools was 38 ng/g lwt. BDE 47 was the most prominent congener, contributing nearly 50% 

of the total concentration, with BDE 99 second at about 20%, BDE 100 at 11%, and so on. 

BDE 209 was present at equal levels in the teenagers living near and working at the disposal site, 

and all other groups, at about 5 ng/g lwt.

In summary, the review of literature on PBDEs in blood has revealed these trends: 

1. Total PBDEs in the general population in the United States generally were in the range of 

30−100 ng/g lwt. The most robust study of U.S adults and older children was based upon 

NHANES 2003-2004 data (2,040 serum samples, all greater than 12 years of age), and 

reported geometric mean concentrations of key congeners BDE 28, BDE 47, BDE 99, 

BDE 100, BDE 153, and BDE 154 at 1.2, 20.5, 5.0, 3.9, 5.7, and 2.3 ng/g lwt, 

respectively (Sjodin et al, 2008). 

2. U.S. levels were higher than levels found in nearly every other study done outside of the 

United States, with most non-U.S. data suggesting total PBDEs to be less than 

10 ng/g lwt total PBDEs.

3. Studies in both the US and abroad showed higher concentrations in toddlers (less than 4 

years old) compared to older ages of children and adults. In paired sampling studies, 

including mother and child, the child’s body burden exceeded the mother’s by about a 

factor of four, exceeding 100 ng/g lwt total PBDEs in some cases. Higher levels in 

toddlers were attributed to ingestion of mother’s milk and high exposures to house dust. 

 

4. The predominant congener in the US and abroad was BDE 47, which comprised about 

50% of the total concentration. The second most commonly found congeners were 

BDE 99 and BDE 153, both explaining about 10−20% of total concentrations. Most 

studies did not measure BDE 209, but, when measured, BDE 209 was found at low levels 

near 5 ng/g lwt. Low levels of BDE 209 have been attributed to the rapid half-life of 

under 10 days in humans. 

5.2.2. Human Milk Data 

 While blood data suggested concentrations of total PBDEs in the range of 

30−100 ng/g lwt, data on PBDEs in human milk suggest possibly higher concentrations, with 

medians or means in some studies in the United States above 100 ng/g lwt. Table 5-2 shows 

congener-specific milk concentrations of BDEs from studies in the United States. Interestingly, 

in one blood study, analysis of results by individuals suggested that females could have 

meaningfully higher concentrations than males. As described above, Schecter et al. (2005) found 

that, in 39 samples from 22 males and 17 females, the range and mean in males were
5-24

Table 5-2. Breast milk concentrations of PBDE congeners in Americans

Congener Concentration

(ng/g lwt) Comment Citation

MonoBDE

3 0.15 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

DiBDE

7 0.01 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

8/11 0.02 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

12 0.01 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

13 0.01 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

15 1.46 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

1.5 WHO interlab calibration study with results from 

17 of 24 countries; United States listed here

Kotz et al. (2005)

TriBDE

17 0.1 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

0.02 (0.01) Mean (median), 47 women in Texas 2002 Schecter et al. (2003)

0.02 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

DL (0.01) N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

25 0.02 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

28 7.4 Mean of 3 pooled sample analyses from CO, CA, 

and NC

Focant et al. (2004)

8.6 WHO interlab calibration study with results from 

17 of 24 countries; United States listed here

Kotz et al. (2005)

2.4 (1.2) Mean (median), 47 women in Texas 2002 Schecter et al. (2003)

1.1 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

1.80, 0.47 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

0.93 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

2 (97%) Median (% detect) from a sample of 303 women 

sampled 3 months post partum in NC

Daniels et al. (2009)
5-25

Table 5-2. Breast milk concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

28/32 3.8 Mean of 40 samples from primaparae women from 

Pacific Northwest

NEW (2004)

28/33 6.2 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

32 0.2 Mean of 40 samples from primaparae women from 

Pacific Northwest

NEW (2004)

35 0.02 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

37 0.06 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

TetraBDE

47 193.0 Mean of 3 pooled sample analyses from CO, CA, 

and NC

Focant et al. (2004)

84.9 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

50; 26 Mean; median of 40 samples from primaparae 

women from Pacific Northwest

NEW (2004)

233.9 WHO interlab calibration study with results from 

17 of 24 countries; United States listed here

Kotz et al. (2005)

35 Partial data from longitudinal study in

Pennsylvania; 3 women

Sjodin et al. (2005)

40.8; 18.4 Mean, median, 47 women in Texas 2002 Schecter et al. (2003)

17.1 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

40.7, 7.7 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

13.9 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

28 (99%) Median (% detect) from a sample of 303 women 

sampled 3 months post partum in NC

Daniels et al. (2009)

49 0.6 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

66 1.1 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

0.3 Mean of 40 samples from primaparae women from 

Pacific Northwest

NEW (2004)

0.65 (0.14) Mean (median), 47 women in Texas 2002 Schecter et al. (2003)

0.07 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)
5-26

Table 5-2. Breast milk concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

1.07, <0.84 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

0.12 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

ND (46%) Median (% detect) from a sample of 303 women 

sampled 3 months post partum in NC

Daniels et al. (2009)

71 0.05 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

0.5, 0.2 Mean, median of 40 samples from primaparae 

women from Pacific Northwest

NEW (2004)

75 0.07 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

77 0.01 (NA) Mean (median), 47 women in Texas 2002 Schecter et al. (2003)

0.02, <0.84 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

PentaBDE

85 5.5 Mean of 3 pooled sample analyses from CO, CA, 

and NC

Focant et al. (2004)

2.3 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

1.1, 0.6 Mean, median of 40 samples from primaparae 

women from Pacific Northwest

NEW (2004)

1.15, 0.41 Mean, median, 47 women in Texas 2002 Schecter et al. (2003)

0.4 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c

1.02, <0.46 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

0.26 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

0.5 (65%) Median (% detect) from a sample of 303 women 

sampled 3 months post partum in NC

Daniels et al. (2009)

99 55.0 Mean of 3 pooled sample analyses from CO, CA, 

and NC

Focant et al. (2004)

20.9 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

10, 5.4 Mean, median of 40 samples from primaparae 

women from Pacific Northwest

NEW (2004)

58.7 WHO interlab calibration study with results from 

17 of 24 countries; United States listed here

Kotz et al. (2005)
5-27

Table 5-2. Breast milk concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

8 Partial data from longitudinal study in 

Pennsylvania; 3 women

Sjodin et al. (2005)

14.0, 5.7 Mean, median, 47 women in Texas 2002 Schecter et al. (2003)

3.5 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

11.8, 1.5 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

2.42 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

5 (92%) Median (% detect) from a sample of 303 women 

sampled 3 months post partum in NC

Daniels et al. (2009)

100 34.4 Mean of 3 pooled sample analyses from CO, CA, 

and NC

Focant et al. (2004)

18.4 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

12, 5.2 Mean, median of 40 samples from primaparae 

women from Pacific Northwest

NEW (2004)

39.2 WHO interlab calibration study with results from 

17 of 24 countries; United States listed here

Kotz et al. (2005)

8.2, 2.9 Mean, median, 47 women in Texas 2002 Schecter et al. (2003)

4.4 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

6.9, <0.46 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

2.40 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

5 (92%) Median (% detect) from a sample of 303 women 

sampled 3 months post partum in NC

Daniels et al. (2009)

118 0.2, <0.46 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

119 0.06 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

126 0.04 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

HexaBDE

138 0.60, 0.09 Mean (median), 47 women in Texas 2002 Schecter et al. (2003)

0.04 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

0.16, <0.03 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)
5-28

Table 5-2. Breast milk concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

0.03 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

138+166 0.26 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

140 0.21 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

153 16.4 Mean of 3 pooled sample analyses from CO, CA, 

and NC

Focant et al. (2004)

19.8 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

16, 4.8 Mean, median, of 40 samples from primaparae 

women from Pacific Northwest

NEW (2004)

17.6 WHO interlab calibration study with results from 

17 of 24 countries; United States listed here

Kotz et al. (2005)

5.3, 2.0 Mean, median, 47 women in Texas 2002 Schecter et al. (2003)

7.0 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

5.2, 1.1 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

3.05 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

6 (99%) Median (% detect) from a sample of 303 women 

sampled 3 months post partum in NC

Daniels et al. (2009)

154 3.6 Mean of 3 pooled sample analyses from CO, CA, 

and NC

Focant et al. (2004)

1.5 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

0.8, 0.4 Mean, median, of 40 samples from primaparae 

women from Pacific Northwest

NEW (2004)

3.0 WHO interlab calibration study with results from 

17 of 24 countries; United States listed here

Kotz et al. (2005)

0.76, 0.22 Mean, median, 47 women in Texas 2002 Schecter et al. (2003)

0.30 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

0.63, 0.05 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

0.17 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

0.3 (54%) Median (% detect) from a sample of 303 women 

sampled 3 months post partum in NC

Daniels et al. (2009)
5-29

Table 5-2. Breast milk concentrations of PBDE congeners in Americans 

(continued)

Congener Concentration

(ng/g lwt) Comment Citation

155 0.3 Mean from 20 primaparae women from around

United States

Lunder and Sharp 

(2004)

HeptaBDE

183 0.15 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

0.3, 0.2 Mean, median, of 40 samples from primaparae 

women from Pacific Northwest

NEW (2004)

0.13, 0.07 Mean, median, 47 women in Texas 2002 Schecter et al. (2003)

0.09 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

6.2, 4.1 N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

0.065 N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)

ND (12%) Median (% detect) from a sample of 303 women 

sampled 3 months post partum in NC

Daniels et al. (2009)

190 0.01 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

NonaBDE

203 <7 (ND) N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

206 0.01 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

207 0.06 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

208 0.01 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

DecaBDE

209 0.24 Mean from 20 primaparae women from around 

United States

Lunder and Sharp 

(2004)

0.8, 0.4 Mean, median, of 40 samples from primaparae 

women from Pacific Northwest

NEW (2004)

0.92, ND Mean, median, 47 women in Texas 2002 Schecter et al. (2003)

0.42 N = 11; mean from blood/milk study (ND = 0) Schecter et al. (2006c)

<204 (ND) N = 38; mean, median from 23 towns in MA Johnson-Restrepo et al. 

(2007)

DL (0.25) N = 46; median from 46 women in the Boston, MA 

area

Wu et al. (2007)
5-30

4.6−192.8 ng/g lwt and 25.1 ng/g lwt, and, for females, the range and mean were 

5.6−365.5 ng/g lwt and 74.1 ng/g lwt. However, Schecter et al. (2005) does note that these 

differences are not statistically significant. 

 The EWG sampled 20 primaparae women from around the country for 35 PBDEs 

(Lunder and Sharp, 2004). Total BDEs averaged 159 ng/g lwt, ranging from 9.5 to 

1,078 ng/g lwt, with six having levels above 100 ng/g lwt and two exceeding 700 ng/g lwt. 

These results were the highest found to date and significantly higher than for European women. 

The most common BDE was BDE 47, which accounted for about half of the total PBDEs in each 

participant. The variability in PBDE levels could not be explained by the diet, occupation, age, 

body mass, or the amount of time they had breastfed their infants. 

 Another environmental organization, the Northwest Environment Watch (NEW), 

conducted a study in 2003 (NEW, 2004). Between April and November of 2003, 40 first time 

breastfeeding mothers from the Pacific Northwest, 10 each from Washington, Oregon, British 

Columbia, and Montana were sampled for the presence of BDE 32, BDE 28/32, BDE 47, 

BDE 66, BDE 71, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209. 

Levels ranged from 6 to 321 ng/g lwt in the human milk, with median and mean levels of 50 and 

97 ng/g lwt, respectively. These levels are comparable to blood and other measurements in the 

United States, but 20 to 40 times higher than levels measured in human milk from Sweden and 

Japan, according to the authors. BDE 47 was the highest found, with a median and mean level of 

26 and 50 ng/g lwt, respectively, with BDE 153 at 4.8 and 16 ng/g lwt, BDE 100 at 5.2 and 

12 ng/g lwt, and BDE 99 at 5.4 and 10 ng/g lwt. BDE 209 was found in 24 of 40 samples; it had 

a high concentration of 4 ng/g lwt and median and mean concentrations of 0.4 and 0.8 ng/g lwt, 

respectively. 

Daniels et al (2009) obtained samples from women participating in the Pregnancy 

Infection and Nutrition studies which were conducted in North Carolina between 2001 and 2005. 

Between 2004 and 2006, a subset of the 589 eligible women remaining in the study were 

sampled twice: months post partum. Nine BDE congeners (BDE 28, BDE 47, BDE 66, BDE 85, 

BDE 99, BDE 100, BDE 153, BDE 154, and BDE183) were measured, and summary statistics 

were provided for these congeners for 303 of the 304 sampled at 3 months post partum (one 

sample was discarded). BDE 47 dominated the profile, with a median of 28 ng/g lwt. Other 

congeners had medians at 6 ng/g lwt (BDE 153, which was the second highest) or lower. Total 
5-31

BDEs equaled the sum of the 5 congeners detected 70% or more of the time (BDE 28, BDE 47, 

BDE 99, BDE 100, and BDE 153), and the median total was 51 ng/g lwt, with a range of 1.2 to 

2,010 ng/g lwt. Looking at the samples from the 83 women providing samples at 3 and 

12 months post partum, there was a wide range of change in total BDE concentrations, with the 

median showing an increase of 14%, an inter-quartile range of −26 to +50%. BDE 153 increased 

the most between sampling dates, and the authors could not provide an explanation for this 

finding. 

 Schecter et al. (2003) collected milk from 47 volunteer donors between August and 

December 2002. Twenty-four donors were from Austin, TX, and 23 were from Dallas, TX. 

Samples were measured for 13 BDEs including BDE 17, BDE 28, BDE 47, BDE 66, BDE 77, 

BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154, BDE 183, and BDE 209. No 

information was available as to whether these women were primaparae. It is, however, 

reasonable to assume they were randomly selected because milk banks were the source of milk 

in Austin and clinics were the source in Dallas. The mean total BDE was 74 ng/g lwt 

(median = 34; max = 419). The dominant congener was BDE 47, comprising 54% of this total, 

with BDE 99 comprising 19% and BDE 153 comprising 7%. BDE 209 was quantified in only 

seven samples, with 16 NDs and the remaining samples not measured for BDE 209. Schecter et 

al. (2003) notes a mean concentration for the 23 samples of 0.9 ng/g lwt. Measurements were 

not correlated with age or length of time in nursing. 

 Focant et al. (2004) evaluated two different analytical methods on both blood and milk 

samples. They analyzed three pooled samples of mother’s milk: one pool from two mothers in 

Denver, one pool from 10 samples collected in 2003 in California, and the third pool from 

10 individuals in North Carolina, also in 2003. BDE 17, BDE 28, BDE 47, BDE 66, BDE 85, 

BDE 99, BDE 100, BDE 153, and BDE 154 were analyzed. The average total BDE from these 

three was 315 ng/g lwt, with BDE 47 the highest at 193 ng/g lwt (61% of total), followed by 

BDE 99 at 55 ng/g lwt (17%), and BDE 100 at 34 ng/g lwt (11%). Results from the more 

traditional gas chromatography and isotope dilute high resolution mass spectrometry 

(GC-IDHRMS) analysis were compared with the newer GC/GC-IDTOFMS, revealing 

reasonable agreement. Because often the number of samples (derived from the different pools) 

analyzed by GC-IDHRMS was higher than the newer method, the summary results are from 

these analyses. 
5-32

Two studies have been conducted in the Boston, MA area. In one, milk was collected 

from 46 women, with total BDE concentrations ranging from 4 to 263 ng/g lwt, with a median of 

28 ng/g lwt (Wu et al., 2007). Only one sample, the highest at 263 ng/g lwt, was higher than 

about 130 ng/g lwt. BDE 47 dominated most samples, although BDE 153 predominated in 

3 samples, and BDE 209 was above detection limits in 11 samples. Questionnaire data 

suggested that concentrations in house dust as well as the consumption of frozen dairy products 

provided the strongest associations with log-transformed total PBDE in human milk. The second 

study in the Boston area was conducted by Johnson-Restrepo et al. (2007), who measured milk 

in 38 volunteer donors between June and November 2004. They measured 17 congeners 

including BDE 209, and found a median total concentration of 19.8 ng/g lipid, with a range of 

0.06 to 1910 ng/g lwt and a mean of 75 ng/g lwt. BDE 209 was not detected in any sample, 

although the detection limit appeared high at 204 ng/g lwt. The most abundant congener found 

was BDE 47, explaining about half the concentrations found. 

 A limited set of longitudinal data (i.e., data on changes over time) was available for three

women for BDE 47 and BDE 99 (Sjodin et al., 2005). These data originate from an ongoing 

study at Pennsylvania State University College of Medicine, where a cohort of 30 participants 

who seek prenatal and pediatric care are being enlisted for a longitudinal study of PBDEs, 

pesticides, and PCBs in milk. Contrary to expectations, levels of all of these contaminants 

suggested increases over time. BDE 47 increased from 30 to 40 ng/g lwt in two of three women 

from Postpartum Day 40 to 120, and from 10 to about 15 ng/g lwt from Day 40 to 60 in the other 

woman. BDE 99 increased from 50 to100 ng/g lwt in one participant from Day 40 to 120, 

increased slightly from approximately 5 to 6 ng/g lwt from Day 40 to 60 in another, and 

decreased from 8 to about 5 ng/g lwt from Day 40 to 90 in the third participant. A more 

comprehensive longitudinal study did show depuration over the course of up to 85 weeks after 

birth (Hooper et al., 2007). Summary statistics of individual congeners (BDE 32, BDE 28/33, 

BDE 47, BDE 66, BDE 71, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183) 

were not provided, and the box and whisker plots for total concentrations (sum of the congeners) 

said little more than that means appeared to be less than 100 ng/g lwt. However, overall results 

suggested that concentrations of all congeners decreased between 1 and 3% per month over the 

course of the study, over the entire range of total concentrations, which was provided at between 

21 and 1,330 ng/g lwt. 
5-33

Ryan et al. (2006) reported on the sampling of human milk from two locations in North 

America: in Hamilton, Ontario, Canada and in Austin, Texas of the United States. The Canadian 

milk samples included 34 that pertained to 2005, 13 that pertained to 2003, 14 that pertained to 

2002, and 26 that pertained to 1992. Specific congeners were not identified; only total PBDEs 

were reported. The lowest concentrations found were in 1992, with a median of 3.1 ng/g lwt 

total PBDEs. Concentrations rose to a median of 33 ng/g lwt in 2003, with another rise to a 

median of 39 ng/g lwt in 2003, and then a decline to a median 20 ng/g lwt for 2005 samples. 

The Texas samples pertained to years 2004 and 2002, and the median concentrations for those 

2 years was higher than these Canadian samples at medians for 2004 and 2002 of 43 and 

44 ng/g lwt 

 Like the blood data, the mother’s milk data suggest much higher concentrations in 

America compared to European and Asian countries. Extracted milk fat from the 3rd round of 

the World Health Organization (WHO)-coordinated exposure study was evaluated for the 

presence of BDEs (Kotz et al., 2005). As of the writing of the study, samples from 17 different 

locations (of 24 total) were used. Nine BDEs were quantified, including congeners BDE 15, 

BDE 28, BDE 47, BDE 77, BDE 99, BDE 100, BDE 126, BDE 153, and BDE 183. The highest 

level by far was found in a sample from the United States at 373.6 ng/g lwt; the second highest 

level was 10.3 ng/g lwt from a sample from Ireland. The predominant congener was BDE 47, at 

63% in U.S. samples (233 ng/g lwt), followed by BDE 99 at 16% (60 ng/g lwt) and BDE 100 at 

11% (41 ng/g lwt).

Schuhmacher et al. (2007) reported on concentrations of 15 BDE congeners (although the 

specific congeners were not identified except to note that samples were fortified with BDE 28, 

BDE 77, BDE 99, BDE 153, and BDE 183; no information on whether BDE 209 was included). 

Their study included 15 women, sampled in 2002, that lived in an urban area (7−10 km from a 

hazardous waste incinerator; 7 women) and near an industrial zone (8 women) in Catalonia, 

Spain. There did not appear to be a distinction in the two small groups, with means of 2.2 and 

2.5 ng/g lwt in the urban and industrial zones, respectively. This same group of 15 women were 

revisited in 2007, and little change in concentrations were noted (Schulmacher et al, 2009). Now 

the mean for the two groups were 2.4 and 2.7 ng total PBDE/g lwt for the urban and industrial 

groups, respectively. 
5-34

 Thomsen et al. (2005b) sampled milk of 151 women representing the northern, 

southwestern, and eastern parts of Norway. Samples were analyzed for BDE 28, BDE 37, 

BDE 47, BDE 85, BDE 99, BDE 100, BDE 119, BDE 138, BDE 153, BDE 154, and BDE 183. 

The sum of the seven most abundant BDE congeners (BDE 28, BDE 47, BDE 99, BDE 100, 

BDE 153, BDE 154, and BDE 183) ranged from 0.95 to 21.05 ng/g lwt, with a median of 

2.35 ng/g lwt, which is comparable to other European countries. BDE 47 was the most 

abundant; it had a median of about 1.15 ng/g lwt. BDE 153 was the second most abundant: a 

median of about 0.50 ng/g lwt. By use of a multiple linear regression model, it was shown that 

there was a statistically significant positive correlation with age (older women had higher 

concentrations) and a negative correlation with parity (number of children) and education. The 

mean concentrations of total BDEs for women <28, 28−31, and >32 was 3.24, 3.39, and 

5.17 ng/g lwt, respectively. The mean concentration for women with one child was 

4.33 ng/g lwt. Women with more than one child had a lower concentration: 3.64 ng/g lwt. The 

authors assert that this was the first study to show statistically significant correlations with age 

and number of children. 

Gomara et al. (2007) report on the sampling of PBDEs in human umbilical cord serum, 

maternal and paternal serum, placentas, and milk from individuals living in two locations 

(Vallecas and Getafe) in Madrid, Spain. The sampling occurred between October 2003 and May 

2004 and involved 391 individual samples including 113 of maternal serum, 104 of paternal 

serum, 92 of umbilical cord serum, 30 of placenta, and 52 of milk. Fifteen individual congeners 

were measured in all samples, including BDE 209. Milk samples had median concentrations in 

the two locations of 6.1 and 5.5 ng total PBDE/g lwt, which was a bit lower than the blood 

samples, which had medians ranging from 9.7 to 17 ng total PBDE/g lwt in the various blood 

matrices. BDE 209 dominated the milk samples, with medians of 2.8 and 2.9 ng/g lwt (ranging 

as high as 52 ng/g lwt) in the two locations. 

A total of 89 lactating mothers in four towns in Japan provided both serum and milk 

samples for analysis of 13 BDEs, including BDE 15, BDE 28, BDE 47, BDE 99, BDE 100, 

BDE 153, BDE 154, BDE 183, BDE 196, BDE 197, BDE 206, BDE 207, and BDE 209 (Inoue 

et al., 2006). The geometric means for the total amounts of the 13 PDEs in human milk and 

serum were 1.56 and 2.89 ng/g lwt, respectively. BDE 209 was the predominant congener in 

serum, accounting for 38% of the total amount of BDEs, but it was a minor component in milk, 
5-35

accounting for 8%. In milk, BDE 47 and BDE 153 were the major contributors, accounting for 

28% and 23% of total BDEs, respectively. 

Fangstrom et al. (2008) conducted a temporal study on Swedish mother’s milk. Fourteen 

pooled milk samples representing 1980 (116 mothers pooled), 1984/1985 (102 mothers), several 

of the years between 1988 and 2002 (20 mothers), 2003 (15 mothers), and 2004 (20 mothers) 

were sampled for BDE 47, BDE 77, BDE 99, BDE 100, BDE 153, and BDE 209. It was 

impossible to quantify BDE 209 in milk, and the authors suggest this could be due to the short 

half-life of this compound, which has been noted for BDE 209 in serum. From the middle of the 

1990s, the concentrations of the lower-brominated BDE congeners (BDE 47, BDE 99, and 

BDE 100) were decreasing, while BDE 153 appeared to be retaining its levels reached towards 

the latter 1990s. BDE 47 was near or less than 0.5 ng/g lwt prior to 1990, reached levels above 

2.0 ng/g lwt by mid-1990s, and dropped sequentially from 1.8 (2001) to 1.4 (2002) to 1.2 (2003) 

to 0.9 (2004). Very similar trends were seen for BDE 99 and BDE 100, with peaks between 0.5 

and 0.8 in the mid-1990s, dropping to 0.3 ng/g lwt in the early 2000s. Meanwhile, the highest 

concentration of BDE 153 until the 1999 sampling was the BDE 99 sample, at 0.82. Afterwards, 

BDE 153 ranged from 0.7 to 1.3 between 2001 and 2004. This trend of decreasing 

lower-brominated BDEs and increasing higher-brominated BDEs was seen in the United States, 

Norway, the Netherlands, and the Faroe Islands. This might be due to declines in use of the 

pentaBDE formulation and/or the debromination of BDE 209 and other higher-brominated 

BDEs.

Lignell et al. (2009) obtained samples from 335 randomly recruited primiparas who lived 

in Uppsala County, Sweden, and delivered between 1996 and 2006. They measured the samples 

for PCBs, PCDD/Fs, hexabromocyclododecane (HBCD), and PBDEs. While declining trends 

over time were noted for PCBs and PCDD/Fs, the results for PBDEs were inconsistent. BDEs 

47 and 99 declined by 4.2% and 7.6%, respectively, while BDE 153 increased 4.9% each year. 

BDEs 100 and 154 were also measured, and BDE 100 showed little change (decline 0.9% each 

year) and results for BDE 154 were not provided. The sum of the five congeners, the total 

concentration, declined 1.7% each year. The mean and median total concentration for all years 

was 3.5 and 2.9 ng/g lwt, respectively. The mean for BDE 47 over all years was the highest at 

1.9 ng/g lwt, while the means for BDEs 99, 100, and 153 were all near 0.5 ng/g lwt. 
5-36

 In a temporal study of organohalogen compounds, including PCBs, PCDD/Fs, and 

PBDEs, in the milk of German women, Furst (2006) found an increase in the mean concentration 

of PBDEs from a pooled sample (n = 300) in 1992 to a sampling of 79 women in 2002. 

Specifically, the mean total concentration (including BDE 28, BDE 47, BDE 66, BDE 85, 

BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183) increased from 1.87 ng/g lwt to 

3.75 ng/g lwt. This contrasted with the other organohalogen compounds studied, which showed 

reductions during this time frame. 

Ohta et al. (2002) determined the concentration of PBDEs in milk of 12 primaparae 

women at 1 month after delivery in Japan. The PBDEs evaluated included BDE 28, BDE 47, 

BDE 99, BDE 153, and BDE 154. Concentrations ranged between 0.7 and 2.8 ng/g lwt in milk. 

Samples were also taken of numerous food products, including 20 fish samples, spinach, 

potatoes, carrots, pork, beef, and chicken (see Chapter 4 for a summary of the results). 

Questionnaires on food consumption were given to the women, and a strong correlation was 

found between consumption of fish and milk concentration. In a “high” group of fish consumers 

(n = 5), the average milk concentration was 1.7 ng/g lwt, while the concentration in the “low” 

group of fish consumers (n = 3) was 0.8 ng/g lwt.

A similar study measuring PBDE congeners in breast milk and tying them to dietary and 

lifestyle habits occurred in North China in 2006 (Zhu et al., 2009). A total of eight congeners, 

including BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209, 

were measured in 80 primiparas living in Tianjin, China. The total concentration ranged from 

1.7 to 4.5 ng/g lwt. The concentration of BDE 209 was the same in blank samples as measured 

in the breast milk, and hence BDE 209 results were not reported. BDE 47 was the most 

predominant congener found, explaining between 22 and 34% of the concentration, although the 

authors noted that this predominance was less than for measurements found around the world. A 

significant correlation was found between concentrations in human milk and the time the woman 

used electronic appliances during the day (hours/day). There were no findings of note relating 

breast milk to diet.

Toms et al. (2007) measured BDEs in pooled milk samples from Australia. A total of 

157 milk samples were collected between 2002 and 2003, and they were pooled to create 

17 regional samples. Eighteen congeners were measured; BDE 209 was not measured. Total 

PBDE averaged 11.1 ng/g lwt (median = 11.0 ng/g lwt) with a narrow range of 
5-37

6.1−18.7 ng/g lwt. BDE 47 dominated the profile, explaining over 50% of the total, followed by 

BDE 99, BDE 100, and BDE 153―all of which contributed between 10 and 20%. 

 In summary, data suggest concentrations of total BDEs in women’s milk in the United 

States exceeded that in blood, perhaps averaging near 100 ng/g lwt, in contrast to a range more 

like 30−100 ng lwt in blood. Part of this trend could be a gender issue, as one blood study 

(Schecter et al. 2005) suggested higher concentrations in women than in men, although this 

difference was not statistically significant. Similar to blood data, concentrations in milk of 

U.S. women exceeded that of women outside of the United States. Milk concentrations of total 

PBDEs outside of the United States were mostly below 10 ng/g lwt. As for blood, BDE 209 was 

usually not measured, or when measured, not found in most instances. BDE 47 was the 

predominant congener, followed by either BDE 99 or BDE 153. 

 

5.2.3. Adipose Tissue Data

Most of the body burden data originate from either blood or human milk. The very 

limited data on adipose are generally consistent with blood and milk data―U.S. data are higher 

than data from other countries; BDE 47 predominates, and BDE 209 is either not measured or 

not found when measured. 

Johnson-Restrepo et al. (2005) measured BDEs in adipose tissue from 52 individuals who 

were undergoing liposuction during October 2003−October, 2004 in New York City. BDEs 

sampled included BDE 28, BDE 30, BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, 

and some unidentified di, tri, and penta-BDEs. The occurrence of BDE 209 was examined 

qualitatively, but it was not detected. Total BDEs ranged from 17.4 to 9,630 ng/g lwt, with a 

median of 77.3 ng/g lwt and a mean of 399 ng/g lwt. The mean dropped to 141 ng/g lwt when 

two outliers of 9,630 and 4,060 ng/g lwt were dropped. The authors claimed this was 10 to 

100 times higher than concentrations reported for human adipose tissue collected from several 

European countries. The data showed no correlation with age, which is consistent with the 

theory that most exposure was only in recent years. BDE 47 was the predominant congener, 

explaining roughly 33% of the total BDEs (at a mean of 132 ng/g lwt), followed by BDE 153 

(23%; 91.8 ng/g lwt), BDE 99 (18%; 74.4 ng/g lwt), and BDE 100 (17%; 67.7 ng/g lwt). 

 Data on 11 adipose tissue samples from 32 women in the industrial port town of Porto 

Alegre, Brazil, showed a range of 0.73 to 3.69 ng/g lwt, with BDE 47 dominating the profile 
5-38

(median = 0.52 ng/g lwt), followed by BDE 99 (median = 0.34 ng/g lwt), BDE 153 

(median = 0.19 ng/g lwt), and BDE 100 (median = 0.12 ng/g lwt). BDE 153 was found in only 

27% of samples, with a median of 0.07 ng/g lwt (Kalantzi et al., 2005). 

Naert et al. (2006) measured BDE congeners BDE 28, BDE 47, BDE 99, BDE 100, 

BDE 153, BDE 154, and BDE 183 in abdominal adipose tissue from 53 individuals (31 men, 

22 women) who died from natural causes or accidents in Belgium. The mean age of the 

individuals was 53 when they died, and no information on diet or exposures was available. 

Consistent with blood data from Europe, total BDE concentrations ranged between 1.23 and 

57.2 ng/g lwt, with a median of 5.32 ng/g lwt. The most predominant congener was BDE 153, at 

a median of 2.40 ng/g lwt, with all other congeners having medians under 1.00 ng/g lwt. 

Fernandez et al. (2007) reported on the concentrations of 14 BDE congeners (BDE 28, 

BDE 47, BDE 66, BDE 71, BDE 75, BDE 77, BDE 85, BDE 99, BDE 100, BDE 119, BDE 138, 

BDE 153, BDE 154, and BDE 183) in adipose tissue from 20 women who were undergoing 

surgery for malignant and benign diseases in Granada, Spain. Three congeners were never found 

above the quantitation limit: BDE 71, BDE 77, and BDE 119. BDE 47, BDE 99, BDE 100, 

BDE 153, and BDE 183 were present at the highest levels and constituted 96% of total 

concentrations. Of these, BDE 153 contributed about 42%, with BDE 47 and BDE 99 2nd and 3rd

most common at 24% and 10%, respectively. The mean concentration of total BDEs (including 

the 11 found above the quantitation limit) was 3.84 ng/g lwt. 

Medina et al. (2008) describe an analytical method for measurement of PBDEs in human 

breast adipose tissue. Their study focused on development and validation of the method; limited 

information was presented on “real” samples they obtained and measured. They measured BDE 

congeners in 15 samples taken in Valencia, Spain, and found monoBDEs to hexaBDE congeners 

(BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) at concentrations ranging from 0.08 to 

0.23 ng/g whole weight (percent lipid in the breast adipose tissue not supplied). BDE 183 and 

BDE 209 could not be quantified because of their low concentrations, but they were detected in 2 

of 15 samples. 

5.2.4. Selection of Representative Body Burden Profiles

 In Chapter 4 (see Section 4.8), exposure media concentrations were assigned for a select 

group of PBDE congeners in order to conduct the exposure assessment. Specifically, these 
5-39

assigned media concentrations were used to estimate exposure intake dose (see Section 5.4 

below). Also, simple pharmacokinetic exercises were applied to those intakes and used to 

predict body burden levels (see Section 5.5 below). To evaluate the merit of that exercise, those 

body burden predictions were compared with measured body burdens. This section will assign 

representative central tendency body burdens for that purpose. 

As also discussed in Chapter 4 in the context of exposure media concentrations, 

assignment of these representative body burdens should have these characteristics: (1) they 

should originate from the United States, or maybe Canada, as representative of industrial North 

American patterns; (2) they should be representative of background and not occupational 

exposures (where occupational exposures pertain to elevated concentrations found in individuals 

working in facilities producing PBDEs or in electronic recycling facilities; not involved in 

everyday office work); and (3) preferably, they should originate from a single study that has the 

ideal characteristics of having a large sample size, being from a diverse geographic area, and 

containing the appropriate set of congeners. Table 5-3 has those assignments―one set each for 

blood and mother’s milk. 

 The blood data in Figure 5-3 were from Sjodin et al. (2008). Of all the U.S. blood data 

available, this evaluation of NHANES 2003/4 was the most representative of recent, national 

trends. Geometric mean concentrations were provided for BDE 28, BDE 47, BDE 99, BDE 100, 

and BDE 153, but not for BDE 17, BDE 66, BDE 85, BDE 154, and BDE 183, because these 

were only quantified in 40% of the samples. Arithmetic means were not provided in Sjodin et al. 

(2008). The median of BDE 99 was reported as less than detection limit, while the geometric 

mean was reported for BDE 99. There were five studies on PBDEs on U.S. human milk that 

could be used to represent background conditions. These included two by public interest 

research groups―the EWG (Lunder and Sharp, 2004) and the NEW (2004), two conducted in 

the Boston, MA area (Wu et al., 2007; Johnson-Restrepo et al., 2007), and a study of 47 samples 

taken in Texas by Schecter et al. (2003). None of the studies had a large sample size. The EWG 

study included 20 samples from around the United States, so it had broad coverage, but it also 

had the highest concentrations of any U.S. study on human milk. The NEW study represented

several states in the Northwest, had a sample size of 40, and its concentrations track well with 

EWG, but they were consistently lower. The Schecter et al. (2003) data on 47 samples were all 

from Texas. Their data also track with the two studies, but it was the lowest of all the data and 
5-40

Table 5-3. Representative body burden levels of PBDEs in Americans 

Description Blooda

(ng/g lwt)

Mother’s milkb

(ng/g lwt)

BDE 17 NA NA

BDE 28 1.2 1.7

BDE 47 20.5 26.0

BDE 66 NA 0.1

BDE 85 NA 0.6

BDE 99 5.0 5.4

BDE 100 3.9 5.2

BDE 138 NA NA

BDE 153 5.7 4.8

BDE 154 NA 0.4

BDE 183 NA 0.3

BDE 197 NA NA

BDE 206 NA NA

BDE 209 NA 0.4

TOTAL 36.3 44.9

a

Blood data from Sjodin et al. (2008). These were geometric means from NHANES. b

Mother’s milk data from NEW (2004). These were medians from a limited opportunistic sample

NA = not available.

originates from one state only. The two studies from Boston, MA, found the lowest 

concentrations, with median levels at 28 ng/g lwt total (Wu et al., 2007) and 19.8 ng/g lwt 

(Johnson Restrepo et al., 2007). The NEW median data from the Northwest was used to 

characterize human milk; it is displayed in Table 5-3. It was chosen because it represented 

several states, it had as large a sample size as the others, and the concentrations were consistent 

among the U.S. studies other than the EWG study, which had the highest concentrations. Mean 

concentrations were also provided in the NEW study, but they were meaningfully higher than 

median concentrations: the median total PBDE concentration was 50 ng/g lwt (this median 

contained more congeners than the median of 44 ng/g lwt shown in Table 5-3) while the mean 
5-41

total concentration was 97 ng/g lwt. Because geometric means were chosen to represent blood 

concentrations, the most analogous statistical representation of human milk concentrations, 

median concentrations, were chosen.

5.3. STUDIES ON INTAKE, OR EXPOSURE DOSE

Several researchers who measured PBDEs in exposure media then went on to estimate 

the exposure dose associated with that media. Their approach was to associate the media 

concentration with a contact rate. For example, a given concentration in dust times an amount of 

dust ingested per day provides an estimate of the daily dose via dust ingestion. This was done 

most often for studies on house dust and food, but researchers also considered inhalation 

exposures when measuring air. A second approach used to estimate intake dose from house dust 

was to measure a surface loading (not a bulk concentration) on the hands in units of mass/unit 

area, and then use an empirical model to correlate that loading with an intake rate, in units of 

mass/time. Other researchers took a more studied approach to evaluating exposure to PBDEs, 

using models, statistical approaches, and other avenues to provide estimates and insights on 

exposure to PBDEs. Table 5-4 provides estimates of these exposure doses. 

 Johnson-Restrepo and Kannan (2009) conducted a comprehensive intake assessment for 

the United States that was structured very similarly to the analysis in this report, using some of 

the same data. They combined dietary intake factors with food data from Schecter et al. (2006b) 

to derive intakes from food. They conducted their own analysis of indoor air and dust from 

homes in Albany, NY, and combined these results with contact rates to derive estimates for 

inhalation, dust ingestion, and dust dermal contact. They combined breast milk concentrations 

from Johnson-Restropo et al. (2007) with infant breast milk ingestion rates to estimate exposure 

to infants from this pathway. They also estimated intake rates for toddlers, children, and 

teenagers. While they concluded that the dust-related pathways dominated exposures for 

children and adults, as did this study, their intake estimates were lower mainly because their dust 

concentrations, at 1,910 ng total/g dust, were lower than the concentrations used in this 

assessment, 8,275 ng total/g dust. Total intakes from all pathways assessed were, in units of ng 

total/kg-day: 86.4 for infants, 13.3 for toddlers, 5.3 for children, 3.5 for teenagers, and 2.9 for 

adults. 

As discussed in Chapter 4, studies have shown house dust to contain much higher 

concentrations of PBDEs compared to outdoor soils. Subsequently, the “soil ingestion” pathway 
5-42

Table 5-4. Estimates of general population intake or exposure dose of total 

PBDEs provided in the literature

Pathway/units Exposure dose/target 

individual Reference/comment

United States and Canada

Total, from dust 

ingestion/dermal 

contact, 

inhalation, and 

diet, ng/kg-day

86.4 for infants

13.3 for toddlers 

5.3 for children 

3.5 for teenagers 

2.9 for adults

Johnson-Restropo and Kannan (2009). They 

combined air/dust measurements from homes in 

Albany, NY with previously published 

food/breast milk to develop comprehensive 

intake estimates.

Dust ingestion 

plus dust dermal 

contact for 

BDE 209 only; 

ng/kg-day

Mean & 95th percentile:

Child—6.0 and 21.0

Adult―0.66 and 2.4 

Petito Boyce et al. (2009); probabilistic using 

data from several U.S. studies for dust 

concentration distribution; distribution and 

point estimates for other exposure parameters

Dust ingestion, 

ng/d 

120−6000 for children; 

3.3 for adults 

Stapleton et al. (2005); children’s estimate 

pertinent to ages 1−4, assuming 0.02−0.2 g/day 

ingestion rate; adult estimate assumes 0.00056 

g/day

Dust ingestion, 

ng/d 

Child median―1380 

Adult median―154

Stapleton et al. (2008b); measurements on 

hands on 33 individuals; empirical model to 

estimate dust ingestion

Dust ingestion, 

ng/d 

400 for adults Sjodin et al. (2004b); based on median of 

4200 ng/g found in Atlanta vacuum dust 

samples, and 100-mg/day adult ingestion

Dust ingestion, 

ng/d 

147; 228 for adults

580; 910 for toddlers 

Harrad et al. (2008a); dust conc. from TX; 

median & average dust conc.; 50, 200 mg/d 

dust ingestion for adults, toddlers

Diet, Dust, 

Inhalation, ng/d 

155 for adults; 264 for 

toddlers (4 mos−2 yrs); 

227 for fish eater; 2190 

for occupational

Jones-Otazo et al. (2005); 66% of adult 

exposure due to dust; 90% of toddler due to 

dust; air/dust modeling + food concentrations; 

Toronto Canada

Diet, ng/kg-d 0.9−1.5 for 

males/females above 

12 yr; 2.6 for 2−5 yr; 306 

for infant

Schecter et al. (2006); based on market basket 

survey and age-based intake rates; infant was 

for breast feeding 

Diet, ng/kg-d 0.3−0.8 for adults Huwe et al. (2005); lower estimate based on 

“lean meats” and higher on meats of high fat 

content
5-43

Table 5-4. Estimates of general population intake or exposure dose of total 

PBDEs provided in the literature (continued)

Pathway/units Exposure dose/target 

individual Reference/comment

Diet, ng/kg-d <1 for children/adults for 

fish; 0.04−20 for 

children/adults for 

beef/chicken

Luksembourg et al. (2004); used market basket 

survey data on fish, beef, chicken from 

Northern CA 

Catfish, ng/kg-d 0.17−1.05; avg. and 95th

fish consumption of 3.6 

and 21.7 g/day

Staskal et al. (2008); catfish consumption using 

measured concentrations of 43 congeners, 61 

fish

Diet + dust + air, 

ng/day 

67−474 ng/day for 

adults, 227−886 ng/day 

for children

Allen et al. (2007); high and low in range based 

on different dust ingestion rates; total includes 

BDE 209

Total dose, (i.e., 

all pathways), 

ng/kg-d

8.5, 16.0, and 53.6 for 

women

McDonald (2005); mean, median, and 95% 

estimate for total dose based on backward PK 

modeling and body burden data for women

Total dose (i.e., 

all pathways), 

ng/kg-d for 

BDE 209 only

1200 for “mid-range” 

and 340,000 for “upper” 

exposure estimates 

Hays and Pyatt (2006) based on 0.96 and 

33.6 ng/g lwt BDE 209 in blood, and backward 

PK modeling

Europe and Asia

Inhalation and 

diet, ng/d 

20; 9 (inhalation) and 

107; 90 (food) for adults 

Harrad et al. (2004); UK, mean, and median 

provided for inhalation and food 

Dust ingestion 

plus dust dermal 

contact for 

BDE 209 only; 

UK only and 

Europe, 

ng/kg-day 

Mean & 95th percentile:

Child/UK—451 and 

1154 

Child/Europe—2.2 and 

8.5 

Adult/UK—54 and 140 

Adult/Europe—0.26 and 

0.96

Petito Boyce et al. (2009); probabilistic using 

data from several European studies for dust 

concentration distribution; distribution and 

point estimates for other exposure parameters

Dust ingestion 

and inhalation, 

ng/d 

0.9−22 (dust ingested) 

and 2 (inhaled) for 

adults; 12−43 (dust 

ingested) and 0.4 

(inhales) for toddlers

Harrad et al. (2006); UK; average results based 

on dust in 8 homes and air in 32 indoor 

locations; mean concentrations & range of 

contact rates; did not include BDE 209 
5-44

Table 5-4. Estimates of general population intake or exposure dose of total 

PBDEs provided in the literature (continued)

Pathway/units Exposure dose/target 

individual Reference/comment

Dust ingestion, 

ng/d 

<600; <11,000 for adults

<2500; <55,000 for 

toddlers

Harrad et al. (2008a, b); UK; median & average 

dust conc.; 50, 200 mg/d ingestion for adults, 

toddlers; BDE 209 dominates

Inhalation, ng/d 0.2, 0.0005−2.9 Mandalakis et al. (2008); median and range for 

inhalation in cars only

Diet, ng/d 97.3 for adults Bocio et al. (2003); Spain

Diet, ng/d 72 and 63 for adults Schuhmacher et al. (2007); industrial and urban 

areas of Spain

Dust ingestion & 

inhalation, 

ng/kg-d

2.0 and 0.2 for children, 

0.2 and 0.4 for adults 

Gevao et al. (2005); Kuwait

Diet, ng/d 62.5, 48.6, and 149.0 for 

adults

Knutsen et al. (2005); Norway

Diet and breast 

milk, ng/d

51 for adults; 110 for 

infants

Darnerud et al. (2001); Sweden; infant dose 

assumed 4.2 ng/g lwt in breast milk

Diet, ng/kg-d 0.79 for adults Bakker et al. (2008); Netherlands; food 

consumption survey combined with composite 

measurements of 47, 99, 100, 153

Diet, ng/kg-d 5.9, of which BDE 209 is 

4.5 for adults

FSA (2006); UK, based on total diet survey 

food samples + consumption data

Diet, ng/kg-d About 1.2 ng/kg-d for 

age groups >15; 

BDE 209 is 1/4 to 

1/3 total

Akutsu et al. (2008); Japan; based on total diet 

survey + consumption data 

Diet, ng/d 35 for adults Voorspoels et al. (2007); Belgium, based on 

market basket samples + consumption data; 

includes BDE 209 

Fish, species 

specific, ng/kg-d 

0.03 for shrimp, 1.35 for 

sole + flounder, 5.46 for 

eels

Bragigand et al. (2006); France; ingestion rate + 

measured BDE congeners from Seine/Loire 

Estuaries

Fish, ng/d

Nursing, ng/d 

Inhalation, ng/d 

1.7−12.9 for child to 

adult 

48.2 for infants 

2.7−9.2 for child to adult

Meng et al. (2007); China; median values for 

fish intake/inhalation for different ages, 0−1 for 

nursing infant 
5-45

has focused on house dust measurements. Stapleton et al. (2005, 2008b) have conducted the 

most comprehensive dust pathway analyses with their studies on house dust coupled with 

exposure modeling. Using estimates of inadvertent ingestion of dust by young children 

(ages 1−4), 0.02−0.2 g/day, they used their measurements of PBDEs in dust to estimate a total 

ingestion of PBDEs range from 120 to 6,000 ng/day (Stapleton et al., 2005). They listed an adult 

exposure of 3.3 ng/day, but this is based on a low estimate of 0.56 mg/day of dust ingestion. 

This value was found in the U.S. Environmental Protection Agency (EPA)’s Exposure Factors 

Handbook (U.S. EPA, 1997), where Hawley (1985) is cited for using a value of 0.56 mg/day to 

characterize adult exposure to house dust from normal activities in the house (higher exposures 

of over 100 mg/day resulted from “work in the attic”). Stapleton et al. (2008b) later approached 

the dust ingestion pathway from a different angle. They measured PBDEs on hands using sterile 

gauze pads soaked in isopropyl alcohol. They measured PBDEs on the hands of 33 individuals 

residing in Durham, North Carolina, and found a median total load per hand of 130 ng, or when 

normalized to surface area, a surface loading of total PBDEs of 135 pg/cm2. They measured 

13 congeners including BDE 209, which was found in 22 of the 33 samples, with a median total 

load of 25.5 ng, and a high of 270 ng. Using an empirical approach based on contact events per 

day, transfer efficiency, the hand loadings, and the fraction of hand coming in contact with the 

mouth, they estimated median exposures to the adult and child to be 154 ng/day and 

1,380 ng/day. 

In contrast to the low dust ingestion rate of 0.56 mg/day assumed by Stapleton et al. 

(2005) in their earlier estimates of PBDE ingestion via dust, Sjodin et al. (2004b) assumed an 

upper limit ingestion rate of 100 mg/day, and using their median concentration of 4,200 ng/g in 

house dust from Atlanta (range of 530−29,000 ng/g), they suggest that this pathway could add up 

to 400 ng/day of PBDE exposure. Such a total would dwarf estimates of 40 to 100 ng/day from 

food ingestion that they cite from the literature. Harrad et al. (2008a) used “mean” and “high” 

dust ingestion rates of 20 and 50 mg/day to model adult ingestion of PBDEs based on 

measurements from homes in Texas. Using arithmetic mean concentrations from dust samples 

(n = 28), they estimated a “high” total intake of 228 ng/day, which includes BDE 28, BDE 47, 

BDE 99, BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209. They also provided estimates 

for toddlers (6−24 months) assuming 50 mg/day as a “mean” and 200 mg/day as a “high” 

ingestion rate. Again using average dust concentrations, a high toddler intake estimate was 
5-46

91 ng/day. They also provided estimates for toddlers and adults in the United Kingdom, New 

Zealand, and Canada, based on dust sampling from those countries. The most interesting finding 

was that UK dust was substantially higher in BDE 209 compared to all other countries, and 

median intakes of BDE 209 were double those from the United States. They continued their 

work by sampling dust in homes, offices, and cars in the West Midland area in the southern part 

of the United Kingdom, and then estimated intakes for adults and toddlers assuming different 

times in the home (72% of the time), office (23.8%), and cars (4.2%) (Harrad et al., 2008b). 

They found totals comparable to their earlier work, including “median” (meaning they used 

median concentrations and high and low dust ingestion rates) intakes for adults to be less than 

5 ng/day for the sum of tri-hexa BDE congeners and 200−600 ng/day for BDE 209. Median 

toddler intakes were less than 10 for the tri-hexa congeners and between 600 and 2,400 ng/day 

for BDE 209. Using data that originated from Kuwait, Gevao et al. (2005) used standard 

exposure assumptions for dust ingestion for children (100 mg/day) and adults (10 mg/day) and 

found that the mean ingestion of total PBDEs averaged 2.0 ng/day for children, and 0.2 for 

adults. Harrad et al. (2006) used house dust measurements from eight homes in the United 

Kingdom to estimate a possible range of adult dust ingestion exposures from 0.9 to 22 ng/day 

total BDEs, and a range for toddlers of between 12 and 43 ng/day. 

Unlike these previous studies on dust ingestion intakes of PBDEs which relied on point 

estimates for intake rates and dust concentrations, Petito Boyce et al. (2009) used a probabilistic 

exposure framework to estimate ranges of intakes of BDE 209 for young children and adults. 

They used a lognormal distribution for the BDE 209 in dust, which assumed a geometric mean of 

1,325 ng/g and a 95th percentile of 8,495 ng/g. They also assumed a log normal distribution for 

child soil ingestion assuming 45 and 124 mg/day as geometric mean and 95th percentile, 

respectively, and for adult soil ingestion, assumed half that of children. They used a combination 

of point estimates and distributions for other necessary parameters (body weight, fraction of total 

ingestion that is dust, etc) and to model soil dermal contact. The mean and 95th percentile 

exposure to BDE 209 via dust ingestion and dermal contact was 0.66 and 2.4 ng/kg-day for 

adults and 4.8 and 18 ng/kg-day for children. Interestingly, they conducted the same analysis for 

European and United Kingdom conditions. For Europe, other than the United Kingdom, they 

found lower intakes, but for the United Kingdom, they found intakes about 100 times higher for 

both children and adults. While BDE 209 concentrations have generally been found at higher 
5-47

levels in the United Kingdom as compared to the United States, this particular result was driven 

by a finding of two very high concentrations of BDE 209 (both above 500,000 ng/g) found in a 

study of 30 homes in the United Kingdom. These findings were also found to be high in followup sampling, and were retained in the analysis by Petito Boyce et al. (2009). 

Similar to the relationship between outdoor soils and house dust, outdoor air 

concentrations were typically lower than indoor air measurements, and, therefore, literature 

estimates of inhalation exposures to PBDEs have focused on indoor air concentrations. Allen et 

al. (2007) provides the most comprehensive evaluation of indoor air exposures to PBDEs. They 

used their own measurements from 20 urban residences in Boston, MA, measured from January 

to March 2006. They measured 11 congeners, including BDE 17, BDE 28/33, BDE 47, BDE 49, 

BDE 66, BDE 85/155, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 209. They used their 

air measurements, along with data on dust concentrations from Stapleton et al. (2005) and food 

data from Schecter et al. (2006b), and combined these with exposure factors from the Exposure 

Factors Handbook (U.S. EPA, 1997), to estimate adult and child exposures to air, food, and dust. 

They estimated total exposures with “mean” dust ingestion rates of 4.16 mg/day for adults and 

45 mg/day for children, and “high” dust ingestion rates of 100 mg/day for adults and 200 mg/day 

for children. Finally, they estimated BDE 209 exposures separate from non-BDE 209 intakes. 

They estimated total intake (non BDE 209 plus BDE 209) for adults as 67 ng/day using “mean” 

dust ingestion and 474 ng/day using “high” dust ingestion, with BDE 209 explaining 23% 

(“mean”) and 31% (“high”) of these totals. For children (ages 1−5, generally, for assignment of 

exposure factors), total intakes were 227 (“mean”) and 886 ng/day (“high”), with BDE 209 

explaining 30% (“mean”) and 31% (“high”). They found inhalation exposures to all congeners 

to be 13% in the “mean” scenario, but much lower at 2% for “high” dust scenario for adults; the 

influence of air for children was much lower, at less than 2%. BDE 209 tended to be the 

predominant congener for inhalation exposures, generally explaining about 40% of total 

inhalation exposures. 

Using data from passive indoor air in a Canadian study, Wilford et al. (2004) assumed 

standard resting respiration rates and found that the median exposure via inhalation was 

1.9 ng/day for females and 2.0 ng/day for males, which compares to a Canadian estimate of 

44 ng/day by dietary intake. Hazrati and Harrad (2005) used passive air measurements in 

12 homes, 10 offices, and 1 private car, and an inhalation rate of 20 m3

/day, to estimate a mean 
5-48

daily intake via inhalation of 4.3 ng/day in the United Kingdom. Harrad et al. (2006) used other 

air data to conclude that average inhalation intakes were 2 ng/day total PBDEs and less in the 

United Kingdom. Using air concentration data from a study in Kuwait, Gevao et al. (2005) 

estimated inhalation doses of 0.4 ng/day for adults and 0.2 ng/day for children (with inhalation 

rates of 20 m3

/day for adults and 8.3 m3/day for children). However, Meng et al. (2007) 

estimated inhalation exposures to PBDEs using data on outdoor urban air concentrations from 

China. They found median exposures doses to range from 2.7−9.2 ng/day. They note this is 

higher than the estimates provided in Wilford et al. (2004) and Harrad et al. (2006), but they 

claim this is because they included BDE 209 in their estimates unlike the other two studies, and 

this congener dominated the air profiles. Mandalakis et al. (2008) studied air concentrations in 

cars in Greece, and combining their measurements with inhalation rates and time in cars per day, 

they estimated inhalation of total PBDEs while driving, to range between 0.0005 to 2.9 ng/day, 

with a median of 0.2 ng/day. This exposure was dominated by BDE 209, explaining about half 

of all exposure. They also found that, despite the small amount of time in the car, that this 

activity contributed 29% of overall daily inhalation exposure. 

 Most direct dose estimates in the literature pertain to dietary dose, and they were

developed mostly by individuals who also measured food concentrations. Estimates of dose 

originate from the United States and overseas, with U.S. studies summarized here first. Schecter 

et al. (2006b) combined their measured average food concentrations with average food 

consumption rates to calculate average intakes for various age ranges (2−5, 6−11, 12−19, 20-39, 

40-59, and > 60) and for males and females separately. Total PBDE intakes ranged from about 

0.9−1.5 ng/kg body weight/day for males/females above the age of 12 (this is the range of

average intakes, derived as noted above over the 4 age ranges, 12-19, 20-39, 40-59, and > 60). 

For ages 2−5 (males and females), the intake was estimated at 2.7 ng/kg-day, and for ages 6−11 

(males and females), the intake was 1.8 ng/kg-day. Individual intakes between BDE 47 and 

BDE 99 were nearly identical for males and females, ranging from 0.4 to 0.7 ng/kg-day after 

age 12, 0.9−1.5 ng/kg-day for the two earlier age ranges. Huwe et al. (2005) estimated a dietary 

intake of PBDEs from meats for a consumer of “lean meats” (5% lipids) was 0.3 ng/kg-day, 

while a “higher fat meats” consumer had an intake of 0.8 ng/kg-day. The estimates are based on 

an average body weight of 53 kg. Fish, meat, and fowl products were purchased in December

2003 and February 2004 from three different food markets in Sacramento and El Dorado Hills in 
5-49

Northern California (Luksemburg et al., 2004). Using average daily intake by adults and 

children taken from the Exposure Factors Handbook (U.S. EPA, 1997), dose estimates were 

provided with their data. Using the highest and lowest concentrations measured in wild and 

farm-raised fish, the theoretical average daily intakes of PBDEs through fish ingestion ranged 

between 0.1 and 1.0 ng/kg-day in children and between 0.02 and 1.0 ng/kg-day in adults. 

Assuming the highest and lowest concentrations measured in beef and chicken products, 

theoretical average daily intakes of PBDEs through ingestion ranged between 0.4 and 

20 ng/kg-day in children and between 0.4 and 10 ng/kg-day in adults. Staskal et al. (2008) 

estimated the dietary intake of PBDEs from consumption of catfish that were store-bought or 

wild-caught (see Chapter 4 for discussion of the sampling and measurement of PBDEs in these 

catfish). Assuming ingestion rates of 3.6 and 21.7 g/day (which they describe as the mean and 

the 95th percentile rate for consumption of all freshwater finfish in the United States), they 

derived a range of ingestion of 0.17 to 1.05 ng/kg-day. This was based on the mean total 

concentration of PBDEs (sum of 43 congeners including BDE 209) in the sample of 61 catfish. 

Twelve studies were found providing dietary dose estimates in Europe and Asia. Bocio 

et al. (2003) estimated that dietary intake equaled 97.3 ng/day for total PBDEs for adults in 

Spain, based on a total diet survey. This was based on homologue group concentrations; a later 

survey of Spanish foods evaluating 15 individual congeners arrived at an estimate of 38.5 ng/day 

for BDEs (Gomara et al., 2006). Schuhmacher et al. (2007) used the food concentration data of 

Bocio et al. (2003), in combination with food consumption rates for an urban and industrial area 

of Spain to calculate dietary intakes of total BDEs of 72 (urban) and 63 (industrial) ng/day. 

Harrad et al. (2004) measured PBDEs in duplicate diet samples from both vegan and omnivorous 

diets. They estimated a dietary exposure average of 107 ng/day (median = 91 ng/day) using 

consumption data from the survey. Additionally, although they did not estimate an exposure 

intake for the vegans, they provided the omnivorous and vegan concentrations, and it is noted 

that the vegan concentrations were about one-half the omnivorous concentrations. They note this 

discrepancy is not as large as the discrepancy for other POPs, like dioxins, which bioconcentrate 

substantially more in animal fat. They could not explain this trend but noted other literature 

showing significant concentrations in vegetative food products. Knutsen et al. (2005) combined 

concentrations from a market basket survey with a comprehensive food consumption survey to 

estimate a mean daily exposure of 62.5 ng/day, a median exposure of 48.6 ng/day, and a 95% 
5-50

exposure of 149.0 ng/day for Norway. Without fish, these numbers were about one-third as 

much, with the mean at 20.0 ng/day and the median at 19.2 ng/day. When a different survey 

more specific to fish types was used, the median rose to 74.2 ng/day, and then even higher when 

a recommended daily additional intake of cod liver oil was assumed (which is recommended by 

the Norwegian government as a healthy supplement)―median intake rose to 122.9 ng/day. A 

similar combination of a comprehensive market basket consumption survey with composite food 

samples measured for PBDEs was conducted by Bakker et al. (2008) for the Netherlands. They 

found a median dietary intake of 0.79 ng/kg-day, with a 95th percentile of 1.62 ng/kg-day, 

dominated by dairy and fish at 39 and 28%, respectively. An estimate of 51 ng total BDEs/day 

was derived for diet only for the Swedish general population (Darnerud et al., 2001). 

Concentrations in a market basket survey were combined with dietary intakes from fish and fish 

products, meat, dairy, and fats/oils. Fish products contributed about half of the total. Meat, 

dairy, and fats/oils contributed about 15% each. Using mother’s milk concentration of 4.2 ng/g 

lwt, they estimated an infant dose of 110 ng/day. Bragigand et al. (2006) measured six BDE 

congeners (BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) in several fish 

species in two major estuaries in France, the Loire and the Seine. Combining ingestion rates (on 

a 60-kg adult basis) for key species with measured concentrations, they noted total intakes of 

PBDEs to be 0.03 ng/kg-day for shrimp, 1.35 ng/kg-day for sole and flounder, and 

5.46 ng/kg-day for eels. 

BDE 209 was not included in any of these dietary intakes and surveys. Instead, the 

standard suite of BDE congeners (BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) was

included. Four studies were found that included BDE 209 as well: one on fish consumption in 

China, one in Belgium, one in Japan, and one in the United Kingdom. Fish consumption intakes 

were determined for a surveyed group of individuals in China, and the median dose ranged from 

1.7−12.9 ng total BDE/day for several age ranges (Meng et al., 2007). While including 

BDE 209, it was found in only 14 of 390 fish samples upon which the intake estimates were 

based. This could have been due to the high detection limit, 0.1 ng/g wwt, of this congener 

compared to detections limits of 0.001−0.003 ng/g wwt of 10 other congeners measured. The 

one in Belgium did not include individual congener breakouts, but it was stated that BDE 209 

was never found above the limit of quantification in any food sample. The average total adult 

intake was only 35 ng/day, generally low but in line with other European surveys that did not 
5-51

include BDE 209 (Voorspoels et al., 2007). A total dietary survey from Japan included BDE 209 

among 36 congeners measured, and it was found at ng/g lwt levels in “fats and oils” (Akutsu et 

al., 2008). Overall, intake of PBDEs from fish dominated the total dietary intake of PBDEs, 

which averaged about 1.2 ng/kg-day for age groups above 15 years old. Fish concentrations 

were dominated by BDE 47 but had concentrations of eight other congeners. BDE 209 in fats 

and oil explained about 1/4 to 1/3 the total intake for all age groups. Childhood intakes, not 

including formula or breast milk, totaled 2.27 ng/kg-day for ages 1−6 (BDE 209 was 

0.90 ng/kg-day) and 1.56 ng/kg-day for ages 7−14 (BDE 209 was 0.59 ng/kg-day).

The presence or absence of BDE 209 could very well be a major issue for these literature 

estimates of European exposures, as a dietary estimate based on a survey that did include 

BDE 209 arrived at substantially different results. The UK Food Surveillance Agency (FSA, 

2006) recently published results of a food survey including BDE 209 and quite alarmingly found 

BDE 209 at the highest level of all BDEs (see Section 4.7). They estimated exposure doses in 

conjunction with their food concentrations and found average intakes totaling 5.9 ng/kg-day, of 

which 4.5 ng/kg-day was due to BDE 209. These results either are questionable themselves, or 

alternately, throw into question other European surveys on “total” dietary dose of BDEs that 

have not measured BDE 209 in the food. It is noted that, in a study on BDEs in dust in Europe, 

BDE 209 dominated substantially over other congeners (Fabrellas et al., 2005), providing 

support to this finding in food, and suggesting that much of the European literature on exposure 

to BDEs is incomplete because BDE 209 was not considered. 

 Jones-Otazo et al. (2005) used models, in combination with reported food concentrations, 

to provide a comprehensive evaluation of exposure to PBDEs pertinent to the Toronto urban 

environment. A regional, multimedia fate model termed Multi-media Urban Model (MUM-Fate) 

was used to predict outdoor soil, outdoor air, indoor, and residential dust concentrations of 

PBDEs. The results were combined with measured concentrations from food and mother’s milk 

to determine potential exposures of PBDEs in this complex exercise. Exposure scenarios 

included the following: elevated indoor sources, fish eater, occupational exposure (from an 

electronics recycling plant), and exposures experienced by four younger age-ranged individuals 

(0−6 months), toddler (6 months−4 years), child (5−11 years), and teen (12−19 years). They 

modeled “total PBDE,” which included BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154. 

BDE 209 was excluded due to a lack of data in most media. Their results suggested first that 
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100−422 g/day of PBDEs were released into the 470-km2 modeled area; these emission rates 

resulted in modeled air concentrations that matched measured concentrations. Their results 

suggest a range of average daily intake from all sources to be 155 ng/day for the adult to 

1,965 ng/day for the infant (2 to 280 ng/kg body weight [bw]). Nearly 100 ng/day of this adult 

exposure is modeled to come from soil/dust exposures, mainly house dust ingestion (dermal and 

inhalation minimal). For toddlers, over 90% of their daily intake of 264 ng/day comes from dust 

ingestion. Following household dust, “dairy, meat, and eggs” exposure contributed 16% of total 

exposures. Other exposure estimates include 227 ng/day for the fish eater and 2,190 ng/day for 

the occupational exposure. While the finding that house dust contributed the most to human 

exposure for all scenarios, this has to be considered carefully in light of the data input into the 

exercise, particularly the food concentrations. It is noted that their assumption of total 

concentration for dairy, meat, and eggs was 101 pg/g wwt, while Schecter et al. (2006b) found 

that total BDE ranged from 39 ppt wwt to 1,426 ppt wwt in 18 meat samples collected in Texas. 

Therefore, it seems quite possible that the food concentrations were lower than would be 

assumed had the estimates been developed for U.S. conditions. 

 McDonald (2005) developed a dose estimate starting from body burdens and working 

backward using PK modeling. Examining 6 studies that evaluated PBDE body burdens in 

individual women (serum, milk, and adipose tissue studies), McDonald found that the median 

concentration was about 48 ng/g lwt, a mean of 90 ng/g lwt, and a 95% of 302 ng/g lwt. These 

totals were the sum of BDE Congeners BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154. 

McDonald used a simple first-order, single compartment (lipid compartment) PK model to 

determine the dose required to obtain congener-specific body burdens. He assumed 

congener-specific adsorptions ranging from 0.78−0.94. Congener-specific half-lives were based 

on rat data and a correlation between rat and human half-lives, resulting in half-lives of 3 years 

for BDE 47, 5.4 years for BDE 99, 2.9 years for BDE 100, 11.7 years for BDE 153, and 

5.8 years for BDE 154. The estimated total dose of PBDEs (sum of 5 congeners) was 8.5, 16.0, 

and 53.6 ng/kg-day for the median, the mean, and the 95% concentrations, respectively. He went 

further to evaluate this dose level in terms of potential for health impact, based on rat testing. He 

found that rodent-to-human body burden concentrations of concern were <1 for alterations of 

male and female reproductive organs in rats, <10 for neurodevelopmental effects in mice, and 

<100 for neurodevelopmental effects in rats. They also looked at other intake studies and noted 
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that an intake estimate of 0.4−11 ng/kg-day was derived by combining food concentrations and 

food intake rates in another U.S. study on food intakes only. 

 Hays and Pyatt (2006) conducted an evaluation similar to that of McDonald (2005) above 

for calculating the dose of the commercial product decaBDE, which is tantamount to an analysis 

of BDE 209, since this congener comprises about 98% of the commercial product. While their 

analysis focused on infant and child exposures to specific pathways and scenarios, and was based 

on a forward intake dose methodology, they also provided an estimate of “general environmental 

exposures” based on blood levels of BDE 209 using the simple backwards PK modeling 

approach. Based on a limited sampling of individuals in the United States (from Sjodin et al., 

2001, reviewed in Section 5.2.1; 12 individuals, age/sex unknown), they assumed a BDE 209 

concentration in blood of 0.96 ng/g lwt to represent the median exposure and 33.6 ng/g lwt to

represent the upper end (UE) of exposures. They used median/high values for half-life of 6.8 

and 3.0 days, and a median/high mixing volume of 25% and 50% body weight (equal to the 

volume of lipids in an adult). Use of these values results in an “absorbed dose.” To derive their 

“intakes,” the amount to which the individual was exposed prior to absorption, they divided their 

absorbed doses by 0.02 for median and 0.01 for upper end (equal to assuming that 2% and 1% 

were absorbed). These absorptions pertain to all routes of exposure. In summary, their analysis 

included different concentrations, different half-lives, different mixing volumes, and different 

absorption assumptions. In their analysis combining these factors, they found a median intake of 

1,200 ng/kg-day and an upper end intake of 340,000 ng/kg-day for BDE 209. These values can 

be compared to the intakes of BDE 209 derived below for general population exposures, 

calculated in a forward manner. It is found that they are substantially higher than calculated in 

this assessment using forward dose modeling. Without providing detail on the methodologies in 

Section 5.4, the unabsorbed intake of BDE 209 for adults is calculated as about 14 ng/kg-day, 

nearly 2 orders of magnitude lower than the “median” estimate of Hays and Pyatt (2006) (note: 

14 ng/kg-day derived from totals provided in Table 5-6; see section below for more detail on 

derivation of the intakes on this table). This is a substantial difference and could suggest that the 

procedures used below to estimate intakes are underestimating exposures to BDE 209 and maybe 

other congeners. However, other differences could bring the numbers together. The review of 

literature below suggested higher absorption than the 2% assumed for calculation of the median 

intake of 1,200 ng/kg-day. For the modeling below, the absorption from dermal exposure to 
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BDE 209 was 3% and for ingestion of BDE 209 by dust ingestion was 4%. These two pathways 

explain over 95% of total exposure in the estimate of 14 ng/kg-day. Also, the literature review 

led to an assignment of 15 days for the half-life of BDE 209 rather than 6.8 days as assumed by 

Hays and Pyatt (2006). Changing their assumption of absorption to 3.5% (from 2%) and 

half-life to 15 days (from 6.8 days) would change their estimate to 316 ng/kg-day (from 

1,200 ng/kg-day). 

Webster et al. (2005) conducted a Monte Carlo exercise on exposures to BDE 47 using 

concentrations of this congener in air, food, dust, and using traditional exposure factors from the 

Exposure Factor’s Handbook (U.S. EPA, 1997). They derived estimates of exposure dose (in 

ng/kg-day) to infants, young children, and adults. Then, using a simple 1-compartment, 

steady-state approach and assuming a 4-year half-life, they estimated body burdens resulting 

from median doses and compared that to some tissue data. From their Monte Carlo simulations, 

the mean dose to infants, children, and adults were 123.9, 7.7, and 0.9 ng/kg-day, respectively. 

The infant dose was dominated by breast milk ingestion, explaining 95% of dose. The children’s 

dose was dominated by the dust-related exposures of dermal (35%) and dust ingestion (36%), 

with diet being the third most important (28%). For the adult, diet dominated at (63%), while 

dust ingestion (23%) and dermal (11%) comprised a significant portion of the remaining 

exposures. Several uncertainties were identified: absorption fractions, concentrations in 

exposure media, and so on. Based on the median adult dose (0.8 ng/kg-day), the steady-state 

lipid concentration was predicted to be 5 ng/g lwt. Assuming breast milk lipid concentrations 

were similar to overall body lipid concentrations, this was lower than the average BDE 

concentration found in other U.S. studies, which the author claimed to average about 20 ng/g lwt. 

The authors suggested possible issues with exposure dose and PK modeling assumptions (4-year 

half-life too short). 

She et al. (2005) used the ratio of BDE 99 to BDE 47 (BDE 99/47) as a way of 

understanding the possible sources of exposure to humans. The study evaluated data from 

numerous studies that provided data to develop ratios in the penta formulation, in house dust, and 

in mother’s milk, and made observations in each. For the penta formulation, BDE 99 and 

BDE 47 are present at about the same amount, 40% each, with BDE 100 at 7%. The ratio of 

BDE 99 to BDE 47 in U.S. house dust ranged from 0.5 to 2.0, with an average ratio of about 1.0, 

which would be consistent with the penta BDE formulation with use information. This suggests 
5-55

the widespread use of penta in North America. The variation in the ratio could be due to 

debromination and different rates of volatilization between the congeners, but overall, U.S. dust 

appears consistent with expectations. UK dust showed high levels of BDE 209, which is 

consistent with the deca formulation. Mother’s milk, on the other hand, showed a BDE 99 to 

BDE 47 ratio of 0.2 to 0.3 in three studies, with the highest levels in milk to have a ratio more 

like 0.6. Their explanations and discussions focused on these points: (1) BDE 47 is more 

bioaccumulative than BDE 99, so that alone could lead to a downward shift in the ratio (i.e., 

more prominence of the more bioaccumulative BDE 47 in mother’s milk); (2) this 

bioaccumulation trend extends not only to direct dust exposures by the mother, but maybe more 

importantly to exposures from animals and then to exposures from animal food products; (3) this 

latter point suggests that on average, a 0.2−0.3 ratio would suggest a predominant pathway by 

food; and (4) the higher mother’s milk concentrations, when the ratio was more like 0.6, might 

suggest that dust is an important pathway as well (because dust is a direct exposure with one 

bioaccumulation step, while food is a secondary pathway). 

The highest estimates of exposure dose, on a body weight, are for infants via breast milk. 

Schecter et al. (2005) calculated infant intakes to be 307 ng/kg-day, dominated by BDE 47 at 

169 ng/kg-day, followed by BDE 99 at 49 ng/kg-day, and BDE 100 at 36 ng/kg-day. As noted 

above, Webster et al. (2005) calculated an intake of 123.9 ng/kg-day for BDE 47 alone, and 

Jones-Otazo et al. (2005) modeled an intake of 280 ng/kg-day for total BDEs for infants ages 

0−6 months. The lowest intake estimates were for breast-feeding infants in China, where Meng 

et al. (2007) calculated a median intake of total BDEs to be about 6−7 ng/kg-day 

(ages 0−1 year). 

 

In summary, some of the key observations from exposure and exposure dose estimation include: 

1. Only one study included a comprehensive evaluation across pathways and age groups, as 

is done in this report (Johnson-Restrepo and Kannan, 2009). They found intakes of total 

PBDEs ranged from 2.9 ng/kg-day for adults to 86.4 ng/kg-day for infants (via breast 

milk). 

2. On a body weight basis, adult exposures for the dietary pathway alone in the United 

States are estimated in the range of 0.5−2.0 ng/kg-day; 
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3. Although no U.S. studies have estimated inhalation exposures, estimates for the United 

Kingdom, Canada, and one study in Kuwait suggest much lower exposures in the range 

of 0.4−>4.0 ng/day, or approximately 0.006 – >0.06 ng/kg-day ; 

4. Adult exposures to PBDEs in dust were estimated to be higher than exposures from 

dietary or inhalation pathways. Exposure to dust ranged from 3.0 to 6.0 ng/kg-day. 

These estimates are made based on vacuum dust sampling combined with dust ingestion 

rates of 100 mg/day; 

5. Estimates for individual routes suggested that exposure to dust dominated adult exposure. 

The majority of studies looking at multiple pathways concluded that dust exposures may 

dominate. However, there was some uncertainty in the literature on this point. One study 

using Monte Carlo techniques found that adult exposures to BDE 47 were dominated by 

diet, 63%, as compared to dust, 23%, and dermal contact with dust, 11% (this was the 

only study found that looked at dermal impacts). Another study assumed house dust 

ingestion was <1 mg/day, so that this pathway resulted only in a BDE intake of about 

3 ng/day; 

6. Child and toddler exposures were higher than adult on a body weight basis 

(approximately 5 ng/kg-day vs. 2.0 ng/kg-day) and were dominated by dust exposures; 

7. Infant exposures were dominated by breast milk ingestion, with body weight-based 

exposures above 80 ng/kg-day, with one estimate over 300 ng/kg-day. 

 

5.4. ESTIMATES OF BACKGROUND INTAKES OF PBDEs FOR ADULTS

 The procedures to estimate individual intakes of PBDEs for adults were developed in a 

manner similar to that done in the EPA Draft Dioxin Reassessment (U.S. EPA, 2003), with one 

important addition, as will be described shortly. Intakes are a function of contact rates in 

combination with exposure media concentrations. As in the Draft Dioxin Reassessment, intakes 

are defined as the amounts of contaminants crossing the body boundary but not yet absorbed into 

the bloodstream. For this reason, absorption fractions are not applied to inhalation and ingestion 

intakes; they are later applied to estimate body burden in the use of the pharmacokinetic model. 

An absorption fraction was, however, used for the dermal contact pathway, as this will result in 

the estimation of an amount crossing the skin boundary. This amount is considered “absorbed,” 

and no further absorption fraction is required for this dermal dose. Table 4-5 provides the 

exposure media concentrations that were developed in Chapter 4, and Table 5-5 in this chapter 

provides the contact rates, and for dermal exposure, the full dermal contact algorithm. 
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Table 5-5. Exposure pathways and factors for the PBDE intake dose 

estimatea

Exposure factors; units Comment/description Adult Ages 

1−5

Ages 

6−11

Ages 

12−19

Body weight, kg Used for converting ng/day to ng/kg 

body weight/day

70 15 30 58

Soil ingestion, mg/d Central tendency values 50 100 50 50

Soil dermal contact, mg/d Surface area that contacts the skin 

(5,700 cm2/d for adults) × amount soil 

adhering to skin (0.07 mg/cm2

) 

× fraction absorbed through skin 

(0.03); area corresponds to head, 

hands, forearms, lower legs

12 2.2 3.2 11

Inhalation, m3/d Unpublished estimates from recent 

studies at 16.1 m3/day for adults

13.3 7.5 12 14

Fraction indoor Children >12 years and adults assume 

21 hr/d; 19 hr/d for children

0.875 0.792 0.792 0.875

Water ingestion, L/d Estimates from U.S. EPA (1997) still 

considered current

1.4 0.69 0.79 0.97

Milk ingestion, g/d Data from USDA (1995) 175 348 357 308

Dairy ingestion, g/d Data from USDA (1995) 55 103 88 77

Egg ingestion, g/d Data from U.S. EPA (1997) in units of 

g/kg bw/d, assumes bw above

16.8 11.25 12.3 13.9

Beef ingestion, g/d Data from U.S. EPA (1997) in units of 

g/kg bw/d, assumes bw above

49.7 21 33 48.1

Pork ingestion, g/d Data from U.S. EPA (1997) in units of 

g/kg bw/d, assumes bw above

15.4 7.2 10.5 15.7

Poultry ingestion, g/d Data from U.S. EPA (1997) in units of 

g/kg bw/d, assumes bw above

35 16.5 26.1 33.6

Other meat, g/d Data from U.S. EPA (1997) in units of 

g/kg bw/d, assumes bw above

24.5 16.5 20.7 24.4

Freshwater/marine fin fish, 

g/d

Data developed in U.S. EPA (2003), 

based on U.S. EPA (2000)

11.6 3 3.8 4.5

Freshwater/marine shellfish, 

g/d

Data developed in U.S. EPA (2003), 

based on U.S. EPA (2000) 

3.8 1 1.2 1.5

a

All exposure factors and approaches developed in EPA’s Dioxin Reassessment (U.S. EPA, 2003), which relied on 

EPA’s Exposure Factors Handbook (U.S. EPA, 1997) and USDA (1995).

 

 The important addition to the procedures originally laid out for dioxin-like compounds, 

applied here to PBDEs, relates to the importance of the indoor versus the outdoor environment. 

The Dioxin Reassessment used air concentrations from outdoor ambient air measurements for the 

inhalation pathway, and measurements of dioxins in background soils for soil ingestion and soil 
5-58

dermal contact. In the case of dioxins, the primary sources are emissions from combustion 

sources into the open environment, with subsequent accumulation in outdoor soils and, of 

primary importance to dioxin exposure, in the terrestrial and aquatic food chains. In contrast, the 

primary cause for PBDE exposures is their use in commercial products that are part of the indoor 

environment (PC circuitry, foam cushions, etc.), and, as described in Chapter 4, indoor air and 

house dust concentrations of PBDEs are orders of magnitude higher than outdoor air and soil. 

Therefore, the use of outdoor measurements in air and soil does not appear appropriate for 

inhalation and soil/dust pathways for PBDEs. Sjodin et al. (2004b) and Stapleton et al. (2005, 

2008b) recognized the importance of exposure to house dust when calculating dust ingestion 

intakes in the hundreds to thousands of ng/day total PBDEs for adults and children. Their intake 

calculations used “soil ingestion” contact rates, applying them in total to their house dust 

measurements, as though the entire contact from “soil” was, in fact, from “house dust.” In 

contrast to their calculations of dust exposure intakes in the hundreds of ng/day, their diet intakes 

(described in the previous section) were always under 100 ng/day total PBDEs. This shows the 

importance of exposures to house dust and indoor air. 

The approach taken in this assessment is to estimate a weighted average concentration of 

“dust/soil” and air, which, in theory, considers what portion of total soil ingestion/dermal contact 

and inhalation comes from house dust and indoor air. The surrogate used to estimate this portion 

will be “time spent indoors.” The Exposure Factors Handbook (U.S. EPA, 1997) provides tables 

on hours/day spent indoors, and for adults, the recommended number of hours/day is 21, which 

is 87.5% (or, expressed as a fraction, 0.875) of the time. Therefore, a weighted average 

concentration, C[avg], of “soil/dust” and air that will be used in the adult soil ingestion, soil 

dermal contact, and inhalation pathways is C(house dust/indoor air) × 0.875 + C(outdoor 

soil/outdoor air) × 0.125. This approach simplistically assumes that exposures are proportional 

to time indoors versus outdoors, which could be totally incorrect for soil pathways if, in fact, the 

actual exposures to dust/soil all were to take place outdoors. In contrast, this is a reasonable 

approach for inhalation. According to the Exposure Factors Handbook, children under 11 years 

of age spend 19 hours per day indoors, so their fraction indoors will be 0.792. Children ages 12 

and higher spend the adult number of 21 hours per day indoors. 

Table 5-5 provides all of the exposure parameters and a brief description of the pathways. 

Table 5-6 provides the final adult intake estimates for the 14 PBDEs for which environmental 
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Table 5-6. Congener-specific and total adult intake estimates of PBDEs (exposures in units of ng/day)a 

Exposure 

pathways 17 28 47 66 85 99 100 138 153 154 183 197 206 209 Total Fraction

House dust 

ingestion

0.00 0.00 81.26 0.92 4.38 102.92 39.86 7.92 10.67 6.86 2.86 0.08 0.01 104.83 362.55 0.66

House dust 

dermal contact

0.00 0.00 19.50 0.22 1.05 24.70 9.57 1.90 2.56 1.65 0.69 0.02 0.00 25.16 87.01 0.16

Inhalation 0.10 0.32 2.15 0.05 0.04 1.00 0.21 0.00 0.06 0.09 0.00 0.00 0.00 1.45 5.47 0.01

Water 

ingestion

0.01 0.00 0.06 0.00 0.00 0.04 0.01 0.00 0.01 0.00 0.01 0.00 0.00 0.06 0.20 0.00

Milk ingestion 0.00 0.01 0.42 0.00 0.00 0.31 0.05 0.00 0.04 0.02 0.03 0.01 0.12 3.01 4.01 0.01

Dairy 

ingestion

0.00 0.02 0.88 0.00 0.00 0.65 0.11 0.00 0.08 0.03 0.07 0.01 0.24 6.30 8.40 0.02

Egg ingestion 0.00 0.00 0.15 0.00 0.00 0.30 0.12 0.00 0.12 0.07 0.00 0.00 0.08 0.57 1.41 0.00

Beef ingestion 0.00 0.03 0.45 0.00 0.00 0.50 0.05 0.00 0.15 0.04 0.02 0.00 0.00 0.15 1.39 0.00

Pork ingestion 0.00 0.01 0.48 0.00 0.01 0.51 0.08 0.01 0.11 0.05 0.31 0.15 0.01 0.25 1.97 0.00

Poultry 

ingestion

0.00 0.01 0.18 0.00 0.00 0.25 0.04 0.00 0.04 0.04 0.04 0.00 0.04 0.14 0.75 0.00

Other meats 0.00 0.01 0.37 0.00 0.00 0.41 0.06 0.01 0.09 0.04 0.17 0.08 0.01 0.19 1.45 0.00

Fresh/marine 

finfish

0.02 0.12 2.32 0.06 0.01 0.46 0.46 0.01 0.07 0.17 0.01 0.00 0.00 0.09 3.82 0.01

Fresh/marine 

shellfish

0.00 0.00 13.68 0.00 0.00 4.56 3.42 0.00 0.00 0.00 0.00 0.00 0.00 0.00 21.66 0.04

TOTAL 0.13 0.54 121.89 1.25 5.49 136.61 54.02 9.85 13.99 9.05 4.20 0.35 0.52 142.19 500.08

Fraction of 

total

0.00 0.00 0.24 0.00 0.01 0.27 0.11 0.02 0.03 0.02 0.01 0.00 0.00 0.28 1.00

a

The sum of dust ingestion and dust dermal contact rounds to 0.90 rather than 0.89, and similarly, the sum of all other pathways rounds to 0.10 rather than 0.09. 

These rounded fractions, 0.90 and 0.10, were used in the text to describe the difference in dust versus other pathways.
5-60

media concentrations were derived in Chapter 4. The total dose to adults was 500 ng/day, or on 

a body weight basis assuming a 70-kg adult, 7.1 ng/kg-day. The following are key observations 

from Table 5-6, and some accompanying discussion: 

1. Predominance of the Dust Pathways: Given the procedures and parameters in Table 5-5, 

coupled with the media concentrations provided in Table 4-5, it would appear that the 

pathways of dust ingestion and dust dermal contact overwhelm the exposure of adults to 

PBDEs. Dust ingestion and dermal contact together account for 90% of the total 

exposure. The estimate of 363 ng/day of PBDEs via dust ingestion is consistent with 

Sjodin et al. (2004b), who calculated an intake of 400 ng/day assuming a dust ingestion 

rate of 100 mg/day and comparable PBDE dust concentrations as were used in this 

exercise. It is higher than the median estimate of 154 ng/day for adults estimated by 

Stapleton et al. (2008b) based on measurements of PBDEs on hands and then an 

empirical model to calculate intakes. The estimate of the PBDE intake by dust ingestion 

is linearly related to the dust ingestion rate and the PBDE concentration. Therefore, use 

of a lower rate of dust ingestion, such as the 0.56 mg/day used by Stapleton et al. (2005), 

would lead to a much lower ingestion of PBDE by the dust ingestion pathway. Further 

discussions on the uncertainty of this pathway are provided in Section 5.7 below. 

2. Prevalence, or Lack Thereof, of Food Exposures:

Europe and Asia. Studies on house dust concentrations for locations outside of the 

United States tend to show lower concentrations of PBDEs (except for BDE 209; see 

discussions in Chapter 4), and so for Europe and Asia, food very likely dominates overall 

exposure. For the United States, the literature is now beginning to recognize the 

importance of the dust pathway, and much more work on dust exposures has occurred 

since about 2005 (see summaries in Section 5.3). 

 After dust exposures, about 10% of 

exposures are due to inhalation and food/water ingestion. In contrast to this finding, the 

literature has an abundance of dietary surveys of PBDEs and discussions on exposure via 

food consumption. A significant portion of this literature deals with food exposures in 

3. Distribution of Dose Among BDE Congeners: The percentage of total dose attributed to 

BDE 47, BDE 99, and BDE 209 are about equal at 24%, 27%, and 28%, respectively, 

followed by BDE 100 at 11%, for a total of 90% among those four congeners. 

Dust-related exposures dominated for the individual congeners. For BDE 47, BDE 99, 

and BDE 209, dust ingestion and dust dermal contact explained 83%, 93%, and 91%, 

respectively. Exposures to BDE 138, BDE 153, and BDE 154 were all low, at between 

2−3% of total. BDE 183, generally considered to be a marker for the presence of the octa 

PBDE formulation, although it could be present in environmental media as a result of 

debromination of higher-brominated BDEs such as BDE 209, was a small contributor to 

overall dose, at 1%. Although at least some measurements were made of all the 

congeners on Table 5-6, it would appear that from a dose perspective, perhaps BDE 17, 

BDE 28, BDE 66, BDE 197, and BDE 206 can be neglected. However, from a body 

burden perspective, BDE 28 makes up between 2−4% of total body burden, while 

BDE 138, which makes up 2% of the dose, is virtually absent in body tissues. This 
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exemplifies how evaluating both dose and body burden can improve the understanding of 

exposure to this class of compounds.. 

5.5. CONVERTING ADULT INTAKE DOSE TO BODY BURDEN

The intake doses in the previous section will be converted to body burdens in this section. 

Assuming first-order kinetics and that PBDEs accumulate in body lipids, the equation for the 

change in lipid concentrations over time is (see eq 5-1) 

δCBDE/δt = [ (DBDE(t) × ABSBDE)/BL(t) ] − k × CBDE(t) (5-1) 

where 

CBDE = the congener-specific lipid-based concentration over time(ng/g lwt) 

DBDE = the daily dose of BDE (ng/day) 

ABSBDE = the congener-specific and route-specific absorption fraction

BL(t) = the body lipid mass over time (g) 

k = the first-order elimination rate of the congener in the body (day−1). 

As presented here, k is assumed to be a constant, but it too could vary over time. The solution to 

this partial differential equation is: 

 CBDE(t) = CBDE(0) × e

(−kt) + [ (DBDE (t) × ABSBDE) / BL(t) ] × [(1 − e−kt)/k] (5-2) 

where 

CBDE(0) = the initial body burden at time 0. 

Assuming a constant BL and a constant dose over time, the steady state lipid concentration (i.e., 

when t approaches infinity) is easily calculated as: 

CBDE = (DBDE × ABSBDE)/(k × BL) (5-3) 
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 Equation 5-3 is used here to estimate an adult body burden, using the congener-specific 

adult doses provided in Table 5-6. It is assumed that bodies are 25% lipid, leading to a BL value 

of 17,500 g. Estimates of absorption for other POPs (dioxins and PCBs) when ingested on soil 

were in the range of 0.30 to 0.70, and it was noted that dioxins might be more bioavailable on 

house dust as compared to soil (Paustenbach et al., 1997, 2006; ATSDR, 2004). Huwe et al. 

(2008) studied the retention and excretion of BDE congeners administered to male rats in corn 

oil and household dust (National Institutes of Standards and Technology reference material). 

The rats were dosed at a “high” dose of 6 μg/kg-day or a “low” dose of 1 μg/kg-day. They were 

fed this amount for 21 days and then killed 24 hours after the last feeding. Fifteen BDE 

congeners were measured in adipose tissue and liver, and feces collected during the experiment 

were also measured to provide a mass balance. By calculating the amounts “retained” from the 

dose in the body by this mass balance, they could surmise absorption by this amount retained. 

The retention amounts for the high dose group exposed to PBDEs in dust, expressed as a fraction 

of total exposure, ranged from 0.04 for BDE 209 to 0.78 for BDE 100. For key congeners 

BDE 47, BDE 99, BDE 138, BDE 153, BDE 154, and BDE 183, the fractions retained were 

0.69, 0.44, 0.67, 0.73, 0.19, and 0.48, respectively. Results were similar for corn oil and for the 

lower exposure amounts. Although not ideal because it is not human data (although human data 

are understandably rare) and it is not corroborated elsewhere, these values were used in the 

modeling of this study for congener-specific values of acrylonitrile butadiene styrene. 

McDonald (2005) assumed absorption fractions of 0.78 to 0.94 for BDE 47, BDE 99, BDE 100, 

BDE 153, and BDE 154 based on experiments with rats, and the values he used were assumed 

for these congeners for all other pathways in this study, including inhalation and water/food 

ingestion. There was no additional absorption assumed for dermal contact because an absorption 

fraction was already included in the dose estimates for that pathway. McDonald (2005) also 

cited Geyer et al. (2004) to assign half-lives of 2.9 to 11.0 years to this same group of congeners, 

and these values were used here. Thuresson et al. (2006) provided estimates of 0.26 years 

(94 days) for BDE 183 and 0.041 year (15 days) for BDE 209, and these were used here. The 

first-order elimination rates, kBDE, were calculated as 0.693/half-life. For lack of better 

information, average values of absorption and half-life, 0.90 and 6 years, respectively, were

assigned to congeners BDE 28 and BDE 138, two other congeners that are present in exposure 

media and in measurements of blood or milk. BDE Congeners BDE 17, BDE 66, BDE 85, 
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BDE 197, and BDE 206 were neglected in this example as they are rarely found in exposure 

media or body measurements. The final set of doses, pharmacokinetic parameters, predicted 

lipid-based concentrations using these parameters in eq 5-3 above, and then the observed 

concentrations of the nine PBDE congeners are shown in Table 5-7. 

Overall, the total concentration was predicted at 31.0 ng/g lwt, while it was observed at 

36.3 ng/g lwt in blood and 44.1 ng/g lwt in milk. Predictions appear reasonably close to 

measurements for five of the eight congeners (for which there are predictions and at least one 

observation). BDE 209 is compared only to a breast milk concentration from Table 5-7, but as 

noted below, the prediction is much lower than what has been found in blood in NHANES. The 

prediction of BDE 47 at 8.6 ng/g lwt does not appear to match well with the observed 

measurements of 20.5 ng/g lwt in blood and 26.0 ng/g lwt in milk. Conversely, the prediction of 

BDE 99 is higher than both measurements, with a prediction of 12.5 ng/g lwt, while it was found 

at 5.0 and 5.4 ng/g lwt in blood and milk, respectively. Otherwise, for example, BDE 100 was 

predicted at 3.9 ng/g lwt, while it was found at 3.9 and 5.2 ng/g lwt. BDE 153 was predicted at 

3.9 ng/g lwt, while it was found at 5.7 and 4.8 ng/g lwt. 

 Obviously, there are uncertainties with use of this pharmacokinetic model and the 

assignment of parameters; these are discussed below in Section 5.7. Even with uncertainties, one 

can make some valid and important observations about exposure of Americans to PBDEs using 

this framework. For one, indoor exposures of soil/dust ingestion, dermal contact, and inhalation 

dominate total exposures. While this might be a valid observation based on intake estimates 

alone, it is strengthened by this simple pharmacokinetic exercise. If one calculates body burdens 

of these nine congeners based on water/food ingestion alone (the dose is estimated at 45 ng/day 

or 0.6 ng/kg-day with a body weight of 70 kg), the body burden is modeled to be 3.7 ng/g lwt, 

which is much less than the observed medians of 36 and 44 ng/g lwt in blood and human milk. 

With 27.3 ng/g lwt predicted to occur from soil/dust ingestion, dermal contact, and inhalation, 

the suggestion is that 88% or more of the U.S. body burden could be explained by nonfood 

exposures occurring in the indoor environment. 
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Table 5-7. Pharmacokinetic parameters and predicted concentrations of BDEs in adults compared with 

measurements in blood and milk

Exposure 

doses/parameters/results 28 47 99 100 138 153 154 183 209 Total

I. Doses

House dust ingestion, ng/d 0.00 81.25 102.92 39.86 7.92 10.67 6.86 2.86 104.83 357.17

House dust dermal contact, ng/d 0.00 19.50 24.70 9.57 1.90 2.56 1.65 0.69 25.16 85.73

Inhalation, ng/d 0.32 2.15 1.00 0.21 0.00 0.06 0.09 0.00 1.45 5.28

Food & water ingestion, ng/d 0.22 18.98 7.98 4.39 0.03 0.70 0.46 0.66 10.75 44.17

II. Parameters

House dust absorption fraction 0.33 0.69 0.44 0.78 0.67 0.73 0.19 0.48 0.04 ---

Other absorption fraction 0.90 0.94 0.78 0.93 0.90 0.90 0.86 0.90 0.90 ---

Elimination half-life, yrs 6.0 3.0 5.4 2.9 6.0 11.7 5.8 0.26 0.04 ---

III. Results

Predicted concentration, ng/g lwt 0.1 8.6 12.5 3.9 1.3 3.9 0.6 <0.1 <0.1 31.0

Observed blood, ng/g lwt 1.2 20.5 5.0 3.9 NA 5.7 NA NA NA 36.3

Observed milk, ng/g lwt 1.7 26.0 5.4 5.2 NA 4.8 0.4 0.2 0.4 44.1
5-65

As currently structured and parameterized, the model may be inappropriate for the 

modeling of the accumulation of BDE 209. The calculated intake dose is 28% of the total, equal 

essentially to intakes of BDE 47 and BDE 99 (calculated at 24% and 27%, respectively). 

However, because of a rapid half-life in the body (15 days), tissue levels are predicted to be as 

low as 0.05 ng/g lwt. This is numerically close to the observed 0.4 ng/g lwt in human milk, but it 

is also 1/8 as much. Some blood studies other than NHANES quantified BDE 209 in blood, also, 

at around 0.1 ng/g lwt. While NHANES did not quantify BDE 209 in individual blood samples, 

Sjodin et al. (2008) note that the concentration of BDE 209 from pooled NHANES 2001/2002 

samples was 2 ng/g lwt (unpublished data; no further details supplied). PSR (2009) quantified 

BDE 209 in 10 of 20 samples, with a range of 3.9–9.0 ng/g lwt. The observation that this pooled 

blood sample BDE 209 concentration of 2 ng/g lwt, or the PSR (2009) findings, are higher than 

the human milk lipid concentration of 0.4 ng/g lwt is consistent with another key study (Schecter 

et al., 2006c) showing a similar disparity in paired samples between human milk and blood 

(samples of milk and blood from lactating mothers; see discussion above in Section 5.2.1). In 

general, this suggests that the model is underpredicting BDE 209 in the body. This could be due 

to an inappropriate assignment of the elimination half-life, or it could be due to an 

underestimation of BDE 209 intakes. In summary, the literature has shown mostly nondetects 

for BDE 209 when measured, and in the range of 1–10 ng/g lwt when found. In contrast, 

BDE 47 and BDE 99 are most often quantified in 90% or more of survey samples, with 

concentrations ranging higher than 20 ng/g lwt. 

However, to observe that BDE 209 has not been quantified at concentrations comparable 

to that of BDE 47 and BDE 99 in blood lipid may not be fully informative. The presence of 

lower-brominated BDE congeners in blood could have resulted from debromination of BDE 209, 

although there is no evidence that the presence of BDE 47 and BDE 99 (and other 

lower-brominated congeners) is the result of the debromination of BDE 209. Also, it may be 

true that BDE 209 is not as lipophilic as other BDEs. Measuring BDE 209 in extracted blood 

lipids might underestimate its presence in blood because some BDE 209 may be present in the 

unanalyzed portion of a blood sample. 
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5.6. EXPOSURE OF SPECIAL POPULATIONS OF INTEREST TO PBDEs 

This section discusses exposures of special populations to PBDEs and develops intake 

and body burden estimates for some of the populations. Specifically, this section: (1) provides 

evidence for exposure to the fetus (umbilical cord blood and placenta measurements, mainly); 

(2) derives intake estimates and resulting body burden estimates for infants; (3) derives intake 

estimates for children within specific age ranges; (4) discusses intakes of BDE 209 derived for 

EPA’s Voluntary Children’s Chemical Evaluation Program (VCCEP) and by the National 

Academies of Science (NAS) using methods different than used in this study; (5) summarizes 

body burdens pertaining to occupational populations; and (6) discusses an important observation 

that there appears to be a proportion of individuals at the high end of the general population who 

experiences significantly higher exposures than the remaining general population. 

5.6.1. Body Burden Data to Characterize Exposure to the Fetus

 The most definitive study showing exposure of the fetus to PBDEs was a study in which 

early to mid-gestation fetal liver (n = 52) and placental tissue (n = 60) samples were obtained 

from elective abortions between 1998 and 2006 in Montreal, Canada, and measured for BDE 47, 

BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183 (Doucet et al., 2009). A clear 

rise in fetal liver tissue was seen over time―total PBDE concentrations (sum of the seven

congeners) averaged 284 ng/g lwt in 1998 and 1,608 ng/g lwt in 2006. Placental tissue 

concentrations were generally lower than fetal tissue, but mean annual concentrations 

(4−8 samples/year) were over 1,000 ng/g lwt for 3 of the 9 years (data presented only in graphs). 

The total concentrations were dominated by BDE 99 (at 42−45% of total) and BDE 47 (30−36%) 

in both liver and placenta, leading the authors to observe that these percentages were similar to 

the pentaBDE formulation, DE-71. 

A second study included measurements of PBDEs in tissues of infants that had been 

stillborn or that died shortly after birth. Schecter et al. (2007) measured levels in fetal liver 

tissue samples that were obtained from four stillborn infants and seven liveborn infants, all of 

whom died shortly after birth and before any feeding. Thirteen congeners were evaluated, 

although BDE 209 was not found in any samples. The mean total concentration was 

19.5 ng/g lwt (assuming ND = 0). At ND = 1/2 DL, the total went up to 23.1 ng/g lwt, owing 

mainly to an average of 3.5 ng/g lwt (ND = 1/2 DL) for BDE 209. The congener profile was 
5-67

fairly similar to blood and human milk profiles, being dominated by BDE 47 at 55% of the total, 

BDE 99 at 23%, and BDE 100 and BDE 153 at 10% and 6%, respectively. One sample was 

uniquely high, 96 ng/g lwt, while all other samples were under 33 ng/g lwt. The high sample 

was from the infant who lived the longest, 7 days. The authors did not clarify why this infant 

was characterized as having died after 7 days before any feeding could occur. While there are no 

data on the levels of PBDEs in the mothers, these levels are consistent with body burdens in the 

U.S. general population as described earlier (Sjodin et al., 2008). 

 Several studies include measurements of PBDEs in umbilical cord blood serum. 

Herbstman et al. (2007) collected umbilical cord blood samples from 297 singleton births from 

11/26/04 to 3/16/05 from Johns Hopkins Hospital in Baltimore, MD. Although limited 

geographically, this was a large number of samples, and a comparison is made here with 

NHANES results for 2003/2004 for the general adult population (Sjodin et al., 2008). Table 5-8 

shows the comparison, which is of data provided in the literature articles and includes limits of 

detection, percent above limit of detection, and median concentrations for congeners where this 

information is reported in the two articles. There does appear to be a difference, although 

statistical tests were not applied to determine if the difference is statistically significant, because 

the detailed data were unavailable but also because statistical testing may not be appropriate 

since the matrices and study designs are different. But from viewing the table, it is seen that 

percent detected is lower for all congeners in umbilical cord blood as compared to NHANES, 

and the median concentrations are also all lower in umbilical cord blood. For example, BDE 47 

was 19.2 ng/g lwt in NHANES but 13.6 from the umbilical cord blood, BDE 153 was 4.8 in 

NHANES and 2.6 ng/g lwt in umbilical cord blood. Since the same laboratory, the Centers for 

Disease Control and Prevention (CDC) laboratory, performed the analysis, this difference was 

not likely to be due to a difference in analytical capabilities. Again, while limited 

geographically, this umbilical cord blood data suggest that the exposure to fetuses may be less 

than the exposure to mothers. 
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Table 5-8. Comparison of median congener concentrations from NHANES 

and median congener concentrations from a large study of umbilical cord 

blood measurements of PBDEsa

BDE congener number

28 47 85 99 100 153

I. NHANES

DL, ng/g lwt 0.8 1.0 2.4 5.0 1.4 2.2

Percent detected 79 97 23 65 93 93

Median, ng/g lwt 1.1 19.2 <LOD <LOD 3.6 4.8

II. Umbilical Cord Blood

DL, ng/g lwt 1.1 1.3 1.3 2.1 1.0 1.3

Percent detected 30 90 15 47 64 60

Median, ng/g lwt 0.9b 13.6 1.1b 4.3b 2.3 2.6

a

The NHANES data are from Sjodin et al. (2008) and the umbilical cord blood measurements are from 

Herbstman et al. (2007). The detection limits (DL) are described as “maximum LOD” in Sjodin et al. (2008), 

and as the “median LOD” in Herbstman et al. (2007). b

Medians are provided in Herbstman et al. (2007) for these congeners despite the fact that over half were below 

LOD. No explanation for how medians were developed was provided in Herbstman et al. (2007).

Data further supporting the hypothesis that the fetus is less exposed than the mother can 

be found in Takasuga et al. (2006), where nine congeners (BDE 28, BDE 47, BDE 99, BDE 100, 

BDE 153, BDE 183, BDE 197, BDE 207, and BDE 209) were measured in maternal blood, 

umbilical cord serum, placenta, and milk from 10 mothers in Japan. In mother’s blood, median 

congener concentrations ranged from a low of 0.08 ng/g lwt for BDE 183 (found in only two

samples) to a high of 1.7 ng/g lwt for BDE 209. Concentrations in mother’s milk were 

comparable to mother’s blood for BDE 28 through BDE 183, but they were lower in milk for 

BDE 197 (0.32 ng/g lipid in blood versus 0.17 ng/g lipid in milk), BDE 207 (0.45 ng/g lipid in 

blood versus 0.12 ng/g lipid in milk), and BDE 209 (1.7 ng/g lipid in blood versus 0.37 ng/g lipid 

in milk), leading the authors to conclude that the transfer from blood to mother’s milk was lower 

for these higher-brominated congeners. All congeners were quantified in the placenta (with 

BDE 183 again found in only 2 of 10 samples) at concentrations that were about one-quarter to 

one-seventh that of the blood. All concentrations were lower than the detection limit (which was 

not provided) for umbilical cord serum. The lower concentrations in placenta, like the lower 
5-69

concentrations seen in umbilical cord blood above, may indicate lower exposure to the fetus as 

compared to the mother. 

Further evidence showing a difference in maternal and umbilical cord blood for 

higher-brominated congeners was seen in a Swedish study (Guvenuis et al., 2003). Samples of 

maternal blood plasma, umbilical cord blood plasma, and milk were taken in 2000−2001 from 

15 mothers living in Stockholm. They were opportunistic samples from women between 28 and 

38 years old. Ten BDEs were measured, including BDE 17, BDE 28, BDE 47, BDE 66, 

BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183. The median and range for total 

concentrations (sum of congeners measured) from the three matrices, in ng/g lwt, were as 

follows: maternal blood―2.1 and (0.7−8.4); umbilical cord blood―1.7 and (0.5−4.3); breast 

milk―2.1 and (0.6−7.7). BDE 47 was the predominant BDE in all matrices (46−70% in milk, 

31−61% in maternal blood, and 45−94% in umbilical cord blood), followed by BDE 153, 

BDE 99, and BDE 100. BDE 47 concentrations were comparable in maternal blood and 

umbilical cord blood, but the levels of BDE 153, BDE 99, and BDE 100 were higher in maternal 

blood as compared to umbilical cord blood (e.g., BDE 153 at 0.56 ng/g lwt in maternal blood but 

only 0.17 ng/g lwt in umbilical cord blood). The authors suggested this might indicate that the 

higher-brominated congeners do not pass through the placenta to the same extent as do the 

lower-brominated congeners. Also noteworthy in this study is the fact that maternal blood and 

maternal milk samples had very similar concentrations for all congeners. The highest 

brominated congener, BDE 183, which is a hepta congener, possibly did not partition as much 

into milk as blood: the median and range in blood and milk, respectively, are 0.06 

(0.01−0.44) ng/g lwt and 0.01 (<0.01−0.14). However, all other congeners showed similar 

median and ranges. This supports the hypothesis that the BDE congeners partition into lipids of 

the body, at least through the hexa congeners. 

Sampling of umbilical cord and maternal blood, as well as mother’s milk, in Canada 

suggests somewhat lower levels in umbilical cord as compared to maternal blood. Ryan and 

Oostdam (2004) collected small amounts of individual maternal and umbilical cord blood 

samples from the Northwest Territories of Canada that were part of a “Northern Contaminants 

Program,” composited them, and measured the composites for congeners including BDE 28, 

BDE 47, BDE 85, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183. These samples 

pertained to years 1994 to 1999. In 10 maternal composite samples, the mean total PBDE 
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concentration was 23.3 ng/g lwt, with a range of 13.1 to 46.5 ng/g lwt, and the umbilical cord 

blood averaged 12 ng/g lwt (range or individual composite samples not provided). In the 

mother’s blood plasma, BDE 47 comprised about 40% of total, whereas in umbilical cord blood, 

BDE 47 comprised about 77%. 

However, other data show umbilical cord blood concentrations of PBDEs to be similar, if 

not even higher, than maternal blood. Gomara et al. (2007) report on the sampling of PBDEs in 

human umbilical cord serum, maternal and paternal serum, placenta, and milk from individuals 

living in two locations (Vallecas and Getafe) in Madrid, Spain. The sampling occurred between 

October 2003 and May 2004 and involved 391 individual samples including 113 of maternal 

serum, 104 of paternal serum, 92 of umbilical cord serum, 30 of placenta, and 52 of breast milk. 

Fifteen individual congeners were measured in all samples, including BDE 209. The maternal, 

paternal, and umbilical cord serum samples had medians that ranged narrowly between 9.7 and 

12 ng total PBDE/g lwt. BDE 47 was the predominant congener in the serum samples. Gomara 

et al. (2007) found that there were no statistical differences between maternal and paternal blood 

concentrations but did find differences between maternal blood versus umbilical cord blood and 

placental concentrations. Unlike the discussion above on umbilical cord blood and NHANES in 

the United States, Gomara et al. (2007) found comparable if not slightly higher concentrations in 

umbilical cord blood. They found significantly lower concentrations in milk and placenta, as 

compared to the blood samples. For example, they found a range of medians between the two 

locations, between paternal, maternal, and umbilical cord blood of BDE 47 of 2.3 and 

3.3 ng/g lwt, but median concentrations of 0.22 to 0.37 ng/g lwt in placenta (one location) and 

milk (from two locations). Interestingly, BDE 209 was highest in the breast milk samples as

compared to all other matrices: the median concentrations in the two locations were 2.8 and 

2.9 ng/g lwt, compared to concentrations around 1.1 ng/g lwt for maternal/paternal blood and 

placenta, and 2.2 ng/g lwt for umbilical cord blood. PBDE median totals were 5.5 and 6.1 ng/g 

lwt for the two sites and 1.8 ng/g lwt for placentas measured at one site. 

Twelve paired samples of maternal and umbilical cord blood were obtained from a 

hospital in Indianapolis during August−December 2001 and analyzed for BDE 47, BDE 99, 

BDE 100, BDE 153, BDE 154, and BDE 183 (Mazdai et al., 2003). Results for maternal and 

umbilical cord blood were essentially identical: the range and median of total PBDE for mother’s 

blood were 15 to 480 ng/g lwt and 37 ng/g lwt, respectively, and the corresponding range and 
5-71

median for infant blood were 14 to 460 ng/g lwt and 39 ng/g lwt, respectively. BDE 47 

accounted for 53−64% of total PBDEs; BDE 99, BDE 100, and BDE 153 each contributed 

10−15% or total. BDE 154 and BDE 183 were found rarely and at low levels. There was no age 

or body mass index relationship with total BDEs. 

Fukata et al. (2005) measured 27 congeners including the key ones (BDE 47, BDE 99, 

BDE 100, BDE 153, BDE 154, BDE 183, and BDE 209) in umbilical cord tissue, maternal blood 

serum, and umbilical cord blood serum in Japan. Samples from eight volunteers were obtained 

and split into Pool A and Pool B (maternal blood was only available from Pool B because there 

was insufficient volume). Umbilical cord tissue samples were uniformly lower for all congeners 

as compared to umbilical cord and maternal serums, which were similar to each other. Total 

PBDEs in the two pools were as follows: 5 and 1.7 ng/g lwt in umbilical cord tissues A and B, 

respectively, 35 and 18 ng/g lwt in umbilical cord serums A and B, respectively, and 20 ng/g lwt 

in maternal serum from Pool B. BDE 209 was not detected in umbilical cord tissue, but it was 

found at 23 and 10 ng/g lwt in umbilical cord serum and 10 ng/g lwt in maternal serum. PCBs 

and chlorinated dibenzodioxins and dibenzofurans (CDD/Fs) had a slightly different trend. 

While umbilical cord tissue was always lower than umbilical cord serum (like PBDEs), they 

were close in magnitude and in fact, maternal serum was significantly higher than either 

umbilical cord tissue or umbilical cord serum. 

Bi et al. (2006) measured the concentrations of seven congeners (BDE 28, BDE 47, 

BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183) in 69 total samples including 21 paired 

maternal and fetal serum (for a total of 42 samples) and 27 milk samples (of which 11 were 

within the 21 individuals providing paired blood samples) in China. They showed a correlation 

between maternal and fetal blood samples (r = 0.55, p < 0.05), with a median total concentration 

of 3.9 ng/g lwt (range: 1.5−12) in fetal serum and 4.4 ng/g lwt (1.6−17) in maternal serum. Milk 

concentrations were comparable though higher than either matrix, at a median of 9.0 ng/g lwt 

and a range of 0.27 to 27.0 ng/g lwt. 

One study showed similar concentrations in umbilical cord blood sera and 4 year-old 

children. The samples were obtained from a large cohort of 470 children who provided data up 

until the age of 4 in Menorca Island in Spain (Carrizo et al., 2007). At the time of sample 

measurement, 92 samples were available from a subset of this cohort who were newborn and 

244 from children age 4. The congeners measured included BDE 17, BDE 28, BDE 47, BDE 66, 
5-72

BDE 71, BDE 85, BDE 99, BDE 199, BDE 138, BDE 153, BDE 154, BDE 183, and BDE 190. 

The average total concentration was 6.2 ng/g lwt in the umbilical cord blood sera compared to 

4.3 ng/g lwt in the sera of the 4-year olds. BDE 47 was the most abundant congener found, 

averaging 2.8 ng/g lwt in the umbilical cord blood and 2.8 ng/g lwt in the sera of the 4-year olds, 

followed by BDE 99 at 1.3 and 1.2 ng/g lwt, respectively. 

This section has reviewed studies that demonstrate that the fetus is exposed to 

polybrominated diphenyl ethers passed on by the mother. Concentrations in fetal liver and 

placental tissue samples from early to mid-gestation aborted fetuses showed concentrations in the 

hundreds to thousands of ng/g lwt total PBDEs (seven congeners up to BDE 183; Doucet et al., 

2009). A second study including fetal tissues from stillborns or from infants who died shortly 

after birth before feedings showed concentrations comparable to national blood measurements in 

the United States (Schecter et al., 2007). In some paired studies including both maternal and 

umbilical cord serum, there is a suggestion that the fetus may be less exposed than the mother, as 

lower concentrations of PBDEs in umbilical cord as compared to maternal serum were found 

(Takasuga et al., 2006; Guvenuis et al., 2003; Ryan and Oostdam, 2004). This trend of lower 

concentrations was seen mostly for higher-brominated congeners in two of these studies 

(Takasuga et al., 2006; Guvenuis et al., 2003). However, other studies showed similar, if not 

slightly higher, concentrations in umbilical cord (Gomara et al., 2007; Mazdai et al., 2003; 

Fukata et al., 2005; Bi et al., 2006). One study showed similar concentrations in umbilical cord

blood sera and sera for 4-year olds (Carrizo et al., 2007). In a large-scale study measurement of 

297 umbilical cord blood samples in Baltimore, MD (Herbstman et al., 2007), all congener 

concentrations found were uniformly lower as compared to national U.S. blood measurements 

from NHANES (Sjodin et al., 2008), and the frequency of detection was also lower. Since the 

same U.S. laboratory, the CDC laboratory, analyzed the samples in both major studies, this 

would not be due to different laboratories doing the work. This may be a reflection of a general 

trend (i.e., lower umbilical cord blood concentrations as compared to general population 

concentrations), but it could reflect the fact that the Baltimore study was regional, while 

NHANES was national in scope. 

 
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5.6.2. Impacts to Infants from Consumption of Breast Milk

Using the profile of PBDEs in mother’s milk, the dose to the infant was modeled as 

follows: 

D = C × f × IR (5-4) 

where

D = ingested dose of PBDEs (ng/day) 

C = concentration in milk fat (ng/g lwt) 

f = fraction of fat in breast milk

IR = ingestion rate of breast milk (g whole weight/day) 

The dose term, D, can easily be converted to a body-weight-based dose term by dividing 

by infant body weight, BW. The rate of ingestion of mother’s milk and the fraction of fat in the 

mother’s milk were assumed to be constant over the duration of breast-feeding. Smith (1987) 

reported that studies in Britain and Houston found that the breast milk ingestion rate for 7- to 

8-month-old infants ranged from 677 to 922 mL/day and 723 to 751 mL/day, respectively, and 

that breast milk ingestion rates remain relatively constant over an infant’s life. Smith (1987) also 

assumed that mother’s milk has a 4% fat content. These assumptions were adopted for the 

purposes of the modeling exercise described here: IR = 800 g/day (which assumes 1-L milk 

weighs 1 kg) and f = 0.04. Given the total PBDE concentration of 44.1 ng/g lwt in mother’s milk 

(see Table 5-3), the total dose to infants is 1,411 ng/day. Assuming an average body weight of 

10 kg for an infant during the months of breast-feeding, a dose is calculated as 141 ng/kg-day. 

This is about half the intake of the 307 ng/kg-day estimate by Schecter et al. (2005). 

 Infant impacts to breast milk and children’s body burdens were handled in a different 

manner than adult body burdens. The approach mirrors what was done for dioxins by Lorber 

and Phillips (2002), who modeled the impact of dioxin-like compounds in infants, resulting from 

consumption of breast milk. The procedures in Lorber and Phillips (2002), as applied to PBDEs 

instead of dioxin toxic equivalents (TEQs), include the following: (1) total PBDEs were modeled 

instead of individual congeners; one half-life (variable as noted below) and absorption (constant 

at 0.80) were used to characterize this surrogate measure of exposure; (2) the dynamic solution to 
5-74

eq 5-2 above, shown in eq 5-3, was used to be able to characterize changes in dose, elimination 

half-life, body lipid fractions, and body weight over time, and (3) the elimination half-life for 

total PBDEs in infants will be more rapid than had been assumed for individual congeners for 

adults. 

 Lorber and Phillips (2002) cited the pharmacokinetic modeling work of Kreuzer et al. 

(1997) in their assignment of the overall elimination rate for dioxin TEQs from infancy into 

childhood. Kreuzer et al. (1997) found that the overall elimination rate of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) in infants was driven by the nonmetabolic 

process of fecal elimination. Specifically, the overall elimination half-life of 2,3,7,8-TCDD was 

modeled to be about 0.4 years at birth, compared to 5 years or more at adulthood, because of the 

magnitude of lipid loss in fecal elimination in infants. Since 2,3,7,8-TCDD accumulates in lipids 

(as do the PBDEs), the rapid loss of lipids via fecal elimination results in a comparable rapid loss 

in 2,3,7,8-TCDD. Lorber and Phillips (2002) assumed this initial, rapid half-life at birth would 

rise to a half-life of 5 years by 18 years of age. They verified their approach by showing how 

infant body burdens of TEQs were predicted to be much closer to measurements in the literature 

when assuming this rapid half-life as compared to assuming the longer half-life that would be 

appropriate for adults. Similar data are not available for PBDEs, but given the similar 

lipophilicity of PBDEs and CDD/Fs, this approach seems reasonable. Specifically, this same 

half-life profile used in Lorber and Phillips (2002) will be used here: a rapid half-life at birth 

rising to an overall, representative half-life of 6 years by age 11. Drawing on information in the 

Exposure Factors Handbook (U.S. EPA, 1997), Lorber and Phillips (2002) also assigned 

temporally-varying body lipid contents and body weights. Absorption of PBDEs will be 

assumed to be 80% (absorption fraction = 0.8), and the initial body burden at birth will be 

37 ng/g lwt, similar to the adult total body burden from blood measurements. Derivation of 

intake dose was provided above, and it was 1,411 ng total PBDE/day for 1 year of breast feeding, 

followed by child intakes from 1 year on (see Section 5.6.2 below). Table 5-9 shows the final 

PK parameters, including dose estimates, from birth until age 19. Figure 5-2 shows the final 
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Table 5-9. Pharmacokinetic parameters for modeling the body burden 

impacts to infants via breast feeding, and then to children from food and 

household exposures

Time after 

birth

PBDE half-life 

(yr) 

Body lipid 

fraction 

Body weight, 

(kg) 

Total PBDE 

dose 

(ng/day)

0 0.40 0.14 3.3 1411

1 mo 0.50 0.16 4.3 1411

2 mo 0.60 0.18 4.6 1411

3 mo 0.70 0.20 6 1411

4 mo 0.75 0.22 6.7 1411

5 mo 0.80 0.23 7.4 1411

6 mo 0.85 0.25 7.9 1411

7 mo 0.90 0.25 8.4 1411

8 mo 0.95 0.24 8.8 1411

9 mo 1.00 0.24 9.2 1411

10 mo 1.05 0.23 9.4 1411

11 mo 1.10 0.23 9.8 1411

12 mo 1.15 0.23 11.3 1411

1 yr, 3 mo 1.30 0.22 11.7 709

1 yr, 6 mo 1.50 0.21 12.5 709

1 yr, 9 mo 1.70 0.20 12.9 709

2 yr 2.00 0.20 13.3 709

3 yr 2.50 0.18 15.6 709

4 yr 3.00 0.16 17.6 709

5 yr 3.50 0.15 19.7 709

6−11 yra 4.00−6.00 0.15−0.13 24−41 389

12−19 yra 6.00 0.13−0.15 41−64 484

a

The time increments of calculation for PK modeling were 1 year after age 5; body weight and lipid fractions 

incrementally decreased/increased during that time within the ranges noted. Doses were constant at the values 

noted.
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Figure 5-2. Modeled infant and childhood body burdens of PBDEs.

results of this exercise, including the impact on body burden assuming the overall half-life of 

6 years. 

With the assumption of more rapid elimination earlier in life, the infant body burden rises 

to about 115 ng/g lwt at age 1 and continues to rise to above 200 ng/g lwt through age 5. At this 

point, the concentration begins to drop, ultimately decreasing to 100 ng/g lwt by age 19. If total 

BDEs had an overall half-life of 6 years from birth on, then the body burden would rise to near 

150 ng/g lwt by age 1 and continue to rise to more than 300 ng/g lwt by age 5, only then to 

dissipate slowly to levels near 100 ng/g lwt by age 19. 

 Two other studies in the literature provide comparable approaches to modeling the breast 

milk exposure pathway for children. Toms et al. (2008) modeled a lifetime of exposures and 

resulting body burden concentrations for individuals in Australia, starting from breast milk 

ingestion at birth, followed by childhood and adult exposures, including food, air, and dust 

0

50

100

150

200

250

300

350

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19

PBDE Body Burden, ng/g lwt

Half-life = 6 yrs

Half-life = 6 weeks at birth

Age, years

0

50

100

150

200

250

300

350

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19

PBDE Body Burden, ng/g lwt

Half-life = 6 yrs

Half-life = 6 weeks at birth

Age, years
5-77

pathways. Their breast milk pathway included a similar ingestion rate (788 g/day), a slightly 

different exposure scenario (6 months exclusive breast feeding, followed by breast plus formula 

feeding up to 1 year), the use of measured mother’s milk concentrations (which also equaled the 

infant starting body burden at birth), and the use of 1-month time step to incorporate growth of 

the infant. They modeled infant body burdens using the adult elimination rates, while the 

analysis here also evaluated a more rapid elimination rate based on the mechanics of lipid 

elimination in the infant. Their rate of ingestion of total PBDEs was 290 ng/day, compared to 

1,411 ng/day assumed in this study. Their results, like these here, show a rapid rise in infant 

body burden through about a year to decline thereafter. Infant body burdens were modeled to 

rise to about 29 ng/g lwt at Year 1, while the infant concentration in this assessment was 

modeled to rise to about 115 ng/g lipid. 

 Hays and Pyatt (2006) also modeled a breast milk pathway but for different purposes, and 

they did not include a module to predict infant body burden. Their study mainly looked at 

intakes of infants and children to decaBDE through ingestion of mother’s milk and then through 

contact with consumer products containing decaBDE. They modeled the breast milk pathway 

for a mother occupationally exposed either through working in a facility that produces deca (the 

mother is involved in product formulation), or through working in a facility that disassembles 

electronics. For each of these two scenarios, they modeled a “mid-range” estimate (ME) in 

which they assumed a 3-month breast-feeding scenario, and for a second estimate they called 

“upper end” (UE), they assumed 2 years of breast-feeding. In one industrial scenario, they used 

limited information on decaBDE in blood from occupationally exposed workers and assumed 

mother’s milk lipid concentration of decaBDE was a fraction of that in blood, based on 

information that decaBDE would not partition to mother’s milk from blood in a 1:1 ratio for the 

highly brominated decaBDE. The ME assumption for this partitioning factor was 0.1, and the 

UE assumption was 0.5. In the other scenario, they modeled breast milk concentrations, starting 

from elevated air concentrations in the industrial setting. They used an air:serum ratio to predict 

serum concentrations and then the same factor less than 1.0 to predict breast milk concentrations. 

Their intake calculations assumed ingestion rates of breast milk of 742 g/day for the ME and 

980 g/day for the UE. Therefore, they produced a wide range of intake estimates, considering 

different industrial settings, different blood-to-milk partitioning factors, different times of 

breast-feeding, and different breast milk ingestion rates. The purpose of their assessment was to 
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compare their predicted intakes to a published reference dose (RfD) of 4 mg/kg-day (NAS, 

2000). When modeling breast milk concentrations starting from air concentrations, the intake 

doses were 80,000 ng/day for the ME assumptions and 2,700,000 ng/day for UE, and when 

modeling based on measured occupational blood measurements of BDE 209, their intake doses 

were 14 and 200 ng/day. These compare with 13 ng/day BDE 209 in this study. The latter two 

estimates from Hays and Pyatt (2006) are comparable to the estimate in this study because of 

comparable breast milk ingestion rates and comparable breast milk concentrations. They started 

with occupational blood measurements of 9.9 (UE) and 4.8 (ME) ng/g lwt, and when 

extrapolated to breast milk using their partition factors of 0.5 (applied to 9.9) and 0.1 (applied to 

4.8), they predicted breast milk concentrations that were similar to the background breast milk 

concentration of BDE 209 of 0.4 ng/g lwt used in this study (see Table 5-3). 

 The validity of the predictions of infant body burdens of BDE congeners presented above 

cannot be easily verified because of the lack of data in the literature. However, there are two

studies described above in Section 5.2 on blood levels of infants and younger children. In one, 

data are presented on four individuals within a family in California: the two parents, 35- and 

37-year olds, a 5-year old daughter, and an 18-month old son sampled in September and 

December of 2004. The sum of BDE 47, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 209 

in the parents ranged between 64 and 147 ng/g lwt for the two sampling dates, but the 

concentrations in the children were much higher. The 5 year-old daughter had concentrations of 

237, 239/249 ng/g lwt (the last two were duplicates of the same December sample) of the five

congeners for the September and then December samples. The toddler had the highest 

concentrations of all: 418 and 488/476 for the five congeners. Also of interest were very high 

initial concentrations of BDE 209, which dropped significantly in both the 5-year old and the 

toddler: the 5-year old had concentrations of 143 ng/g lwt of BDE 209 in the September sample 

and 9 and 12 ng/g lwt (duplicate samples) in the December sample. The toddler had 233 and 

19/26 ng/g lwt in the September and December samples, respectively. Discounting laboratory 

error, the authors attribute the higher concentrations in the children to exposure to house dust, 

and the drop in BDE 209 levels between the September and December samples to the short 

half-life of BDE 209 in humans. While the authors have discounted laboratory error, it would 

appear that a decline by an order of magnitude in both the toddler and infant is substantial and 

could be due to some difference in the two laboratories. However, a decline of this magnitude is 
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plausible because of the short half-life of BDE 209 in humans, as noted by the authors. The 

higher levels of the other congeners in the toddler were attributed by Fisher to consumption of 

breast milk, although it could also be due to exposure to house dust. 

The second study was a large-scale pooled blood sampling survey in Australia (Toms et 

al., 2008). As described in Section 5.2, the highest concentration was found in the youngest age 

group pooled: 0−4 years old with a mean total PBDE concentration of 73 ng/g lwt, with 

subsequent lower concentrations for older groups and an average for adults (all age groups above 

16 years old) about 15 ng/g lwt. Their modeling of infant and childhood impacts, which showed 

only about a maximum of 29 ng/g lwt at age 1, was noted by the authors to underpredict the 

observed body burdens of children. A third study included 244 measurements of 4-year old 

children of a large cohort of 470 children studied on the island of Menorca in Spain (Carrizo et 

al., 2007). They found higher concentrations in children who had been breast fed as compared to 

formula fed, and the differences were statistically significant for the congeners found at the 

highest concentrations, BDE 47 and BDE 99. The average total concentration (sum of 13 

congeners not including BDE 209) in the 4-year olds that had been breast fed (n = 202) was 

3.6 ng/g lwt, compared to 1.3 ng/g lwt for formula fed (n = 44). 

 These high measurements in children support the simple PK modeling done here, but are 

not sufficient to verify it. Body burden measurements in children remain an uncertainty, 

although the modeling and these observations provide some evidence that body burdens in very 

young children are higher than adults, with the suggestion that they could be very high due to 

consumption of breast milk. Further, the analysis here suggests that these higher levels could 

persist throughout childhood into early adulthood. 

5.6.3. Childhood Intakes

The total dose to adults was 500 ng/day, or 7.1 ng/kg-day (assuming an adult body 

weight of 70 kg). Using the intake rates provided in Table 5-5 and the exposure media 

concentrations developed in Chapter 4 (shown in Table 4-5), the total dose for the three age 

ranges of children were as follows: 709 ng/day for the 1−5 age range, 389 ng/day for the 

6−11 range, and 484 ng/day for the 12−19 age range. On a body weight basis, the doses are 

47.2 ng/kg-day for ages 1−5 (assuming 15 kg bw), 13.0 ng/kg-day for 6−11 (assuming 

30 kg bw), and 8.3 ng/kg-day for 12−19 (assuming 58 kg bw). The much higher dose for the 
5-80

child age 1−5 was due to the doubling of soil/dust ingestion from 50 mg/day to 100 mg/day. 

Otherwise, the trends as elaborated above for adults in Section 5.4, such as the predominance of 

the soil ingestion and dermal contact pathways, were similar for children. One study suggests 

that these intakes may be low for children. Stapleton et al. (2008b) evaluated PBDEs that are 

present on hands from 33 volunteers using wipe samples, and using a hand-to-mouth algorithm 

for children, they estimated a median intake for children (generally defined as 6−13 years old 

based on parameters selected to model hand-to-mouth intakes) to be 1,390 ng total BDE/day and 

95% of 6,090 ng total BDE/day (including 13 BDE congeners up to BDE 209 that comprised 

roughly 1/5 of total exposures). These numbers were generated using the data from all the 

participants, and only six of them (ages 8 to 10) could have been characterized as children, and 

the authors acknowledge this to be a key uncertainty for these intakes for children. 

5.6.4. Intake Estimates Derived for EPA’s VCCEP and by the NAS

The intake estimates derived for EPA’s VCCEP (described in Hays and Pyatt, 2006) and 

by the National Academies of Science (NAS, 2000) differed from the intake estimates derived in 

this assessment in two ways: (1) the focus was solely on the commercial product, decaBDE, 

which essentially is equal to assessing BDE 209 since this congener comprises about 98% of the 

commercial product, and (2) the purposes of both studies was to provide conservative intake 

estimates with which to compare to the NAS reference dose of 4 mg/kg-day for decaBDE. 

Because of this second difference, many of the intakes provided below were significantly higher 

than intakes estimated in this assessment, by upwards of 5 orders of magnitude. 

The VCCEP intake assessment was presented to EPA as part of a larger assessment of 

decaBDE that included an assessment of toxicity, environmental levels, and many other analyses 

(VCCEP, 2008). Hays and Pyatt (2006) published this VCCEP assessment of exposure to 

children in the open literature. They modeled five exposure scenarios pertaining to children 

ages 0−2: (1) breast milk ingestion from a mother who is involved in the formulation of 

decaBDE, (2) breast milk ingestion from a mother who is dissembling electronics, (3) mouthing 

decaBDE-based plastic electronic products, (4) inhaling decaBDE particulates released from 

plastic electronic products, and (5) mouthing decaBDE-containing fabrics. They modeled a 

6th scenario pertaining to all ages, which was a backward PK modeling exercise starting from a 
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blood measurement of BDE 209. This exercise was described earlier in Section 5.3. The two 

breast milk intake estimates were described in Section 5.6.2. 

 The other three scenarios involving mouthing of products (fabrics and electronics) and 

inhaling particles released from products (electronics) involved estimating the release from the 

products and then the exposure by the children. Their highest exposure was from mouthing 

fabrics, which they used from the NAS (2000) assessment described below. The intake was 

26,000 ng/kg-day of decaBDE for this pathway. The general algorithms for mouthing fabrics 

and mouthing electronics present an alternate to the dust ingestion and dust dermal contact 

algorithms in this assessment, which are the closest analogous algorithms. The general 

algorithm for these mouthing algorithms entails an estimate of the total mass of the PBDE in the 

product per square area, a rate at which it leaves the product or can leave upon mouthing, and 

then exposure contact parameters. Data to parameterize these approaches are sparse, and the 

assignments in Hays and Pyatt (2006) and NAS (2000) can lead to very high intake estimates (as 

will be shown in later discussions). For mouthing electronics, the total intake entails an estimate 

of the mass of decaBDE leaching from the surface of the product (0.15 mg/day for the 

“mid-range” estimate and 0.29 mg/day for the “upper” estimate, as generated from a study in the 

literature), combined with exposure factors including the time spent by the infant mouthing 

during the day (32 and 97.2 minutes/day) and the fraction of items he/she mouths that is treated 

(1% and 10%). Their intakes for BDE for this “mouthing” scenario were 4.3 and 250 ng/kg-day. 

Their inhalation scenario did not entail a release and air dilution algorithm but instead was based 

on measured concentrations of decaBDE in air. Their inhalation intakes were 0.3 and 

0.6 ng/kg-day. In contrast, the dust ingestion, dust dermal contact (not considering absorption 

across the skin barrier), and inhalation intake estimates for BDE 209 for this assessment for the 

youngest age category, 1−5 years, were 7.0, 1.7, and 0.1 ng/kg-day (based on an average 15 kg 

body weight during the Years 1−5). The approach in this assessment was to use background 

concentrations of BDE congeners in air and house dust and combine these contact fractions to air 

and dust. It did not consider more specific contact or behaviors associated with treated products. 

The inhalation estimates in this assessment and the Hays and Pyatt (2006) assessment are 

similar, because the approaches were the same, both using air concentrations combined with 

contact fractions. Their mid-range estimate for mouthing of electronic products led to exposures 
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similar to the dust pathways in this assessment, but their upper estimate was over an order of 

magnitude higher. 

 The NAS study also had different intake estimates for decaBDE as compared to the 

BDE 209 estimates in this study (NAS, 2000). For dermal exposure to treated fabric, their 

general model entails first an estimate of release from the treated fabric, followed by 

consideration of clothing as a barrier to skin transfer, and then an assumption that enough 

moisture (sweat) is present such that decaBDE can be dissolved and transferred through the skin 

barrier. A first iteration of their model assumed that the clothing did not prevent a barrier to 

transfer, and then additionally, full transfer through the skin such that the only limitation to 

exposure was the rate of release from the treated fabric. Under this extreme example, they 

estimated an exposure as 980,000 ng/kg-day. When they modeled penetration through the skin 

(still not considering clothing as a barrier), they estimated an absorbed intake of 

0.001 ng/kg-day. As noted above, the dermal contact pathway in this assessment estimates

BDE 209 exposures of 1.7 ng/kg-day to dust before absorption, and when considering a 4% 

absorption rate, this is reduced to 0.07 ng/kg-day. An algorithm to consider sucking of treated 

fabric includes an area density of treated fabric (mass decaBDE per unit area of cloth), an area of 

sucking, an extraction rate (a potential amount of mass extracted for a full day of sucking), and 

an amount of time sucking. The average dose rate using this algorithm was very high at 

26,000 ng/kg-day, considerably higher than the 7 ng/kg-day estimated for dust ingestion in this 

assessment. An inhalation pathway models both particle and vapor phase decaBDE released 

from treated fabrics. The empirical model is based on area density of treated surfaces, the extent 

of treated surface within a room, a rate of release of the decaBDE as respirable particles, and the 

room volume in order to produce a concentration. Finally, a time-weighted average daily 

concentration considered the amount of time an individual spent in the treated room. The 

time-weighted average was 480,000 ng/m3

. Similar algorithms were used to estimate a vapor 

phase concentration, and it was found to be 380 ng/m3. These modeled concentrations are 

extremely conservative, as BDE 209 has not been found to exceed 1 ng/m3 in the 

nonoccupational setting, and only as high as the hundreds of ng/m3 in occupational settings (see 

Section 4.5). In NAS (2000), these predicted concentrations were compared to a Reference 

Concentration, RfC, determined from their RfD. 
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 There is some concurrence in the approaches taken by Hays and Pyatt (2006) and NAS 

(2000) and this assessment. Such concurrence was seen in inhalation exposures based on 

measured air concentrations, and in the dermal contact pathway when considering the skin as a 

barrier to absorption. However, mostly the intake algorithms and air concentration models led to 

substantially higher intakes and concentrations as compared to estimates in this 

assessment―higher by up to 5 orders of magnitude. Certainly, the objectives of the Hays and 

Pyatt (2006) and NAS (2000) study were different than here. A key purpose of their assessments 

was to compare exposure intakes with the NAS (2000)-derived RfD of 4,000,000 ng/kg-day 

(4 mg/kg-day). There was never any attempt to compare predicted with measured 

concentrations, nor any attempt to compare intake estimates with others that may have been 

generated in the literature on similar pathways. Also, there was no attempt to take intakes and 

model a blood concentration of BDE 209, which could have been used to compare with 

measured blood concentrations (although there were few, if any, measurements of blood 

concentrations of BDE 209 for use in the NAS report completed in 2000). While the algorithms 

used show potential use in assessments of general population exposures, it is felt that the input 

parameters need to be more carefully assigned values to characterize general population 

exposures, and the results subjected to some reality or validation testing. 

5.6.5. Body Burden Data to Characterize Occupational Exposures

Limited studies from Sweden and from China suggest that PBDE concentrations are 

elevated in occupational groups exposed to likely sources of PBDEs. One study which looked at 

incinerator workers in comparison to general population exposures, did not find a difference (Lee 

et al., 2007). The only study of occupational exposures in the United States found a significantly 

higher (p < 0.05) level of PBDEs in workers in foam recycling facilities and individuals who 

installed carpet padding manufactured from recycled foam (Stapleton et al., 2008a). 

The study from the United States included 12 foam workers from two foam recycling 

facilities (one in Maryland and one in California), 3 carpet layers who worked in association with 

the California facility, and 5 control group individuals composed of spouses and clerical workers 

from the facilities (Stapleton et al., 2008a). The median total PBDE concentrations (composed 

of BDE 17, BDE 28, BDE 47, BDE 66, BDE 99, BDE 100, BDE 153, BDE 154, and BDE 183) 

in the foam workers, carpet installers, and control were 160, 178, and 19 ng/g lipid, respectively. 
5-84

The body burdens of the workers were dominated by BDE 47, which explained 50−60% of the 

total concentration, followed by BDE 99 and BDE 153, which both contributed 13−20% of the 

total. 

Thirteen PBDE congeners were quantified in serum in a group of 19 PC technicians (PC 

techs), and the results were compared against hospital workers, and PC clerks in Sweden 

(Jakobsson et al., 2002). Moreover, within these two comparison groups, refined 

characterizations included women who had never breast fed (NBF), hospital cleaners (HCs), and 

PC clerks (PCCs). Results were provided for BDE 47, BDE 153, BDE 154, BDE 183, and 

BDE 209. There were distinct differences among the groups. The median value of BDE 47 was 

fairly similar among the groups: PC techs had a median of 1.3 ng/g lwt, HCs had a median of 

1.6 ng/g lwt, PCCs had a median of 1.5 ng/g lwt, and NBFs had a median of 2.1 ng/g lwt. 

BDE 153 was the highest in the PC techs and significantly higher than the other three groups: 

2.6 ng/g lwt for the PC techs, while it was 0.6, 0.8, and 0.8 ng/g lwt for the other three groups. 

While these other three groups did not have BDE 209, it was quantified at a median of 

1.5 ng/g lwt for the PC techs. There were no correlations with age and BDE levels, but a 

correlation was found between time on the job and BDE 153 concentrations in the PC techs. 

Eight Swedish employees at a recycling plant and four rubber mixers volunteered to 

donate blood samples during their summer vacations, in 1998 and 2000, respectively (Jakobsson 

et al., 2003). The first blood sample was drawn at the start of vacation, 3−4 days later, and then 

at additional time frames during the 4−5 week vacation. The samples were analyzed for 

BDE 47, BDE 153, BDE 183, and BDE 209. Assuming there was a constant baseline of BDEs 

in these workers, which they obtained from serum measurements on nonexposed individuals, the 

authors calculated half-lives. In addition to these 8 temporally followed workers, the authors 

measured blood from 60 other subjects; in all, their study had 107 observations from 68 subjects. 

They found distinct exposure levels and patterns of BDE congeners as a function of which 

workers they measured. The rubber workers, who were exposed only to decaBDE, had markedly 

elevated levels of BDE 209, which averaged 27−35 ng/g lwt, with a high of 278 ng/g lwt. 

Individuals not occupationally exposed had levels at ND−2.4 ng/g lwt. The electronic 

dismantlers were expected to be exposed to all BDE congeners, and they were, with 

concentrations of BDE 47, BDE 153, BDE 183, and BDE 209 averaging 2.9, 4.5, 7.9, and 

4.8 ng/g lwt, respectively. While they found that BDE 209 is bioavailable based on finding high 
5-85

levels in an occupationally-exposed group (the rubber workers), they also found it had the 

shortest half-life. There was, overall, an inverse relationship between half-life and degree of 

bromination―lower-brominated congeners had longer half-lives: BDE 203 was 37 days, 

BDE 183 was 110 days, BDE 153 was 680 days, and BDE 154 was 270 days. 

 Qu et al. (2007) measured levels of 14 BDEs (including BDE 209) in electronic waste 

dismantling workers, in residents living within 50 km of the dismantling area, and a reference 

group with no exposure near an occupational setting in Quongdong, South China. The median 

concentrations of each congener were determined, and the sum of these individual medians from 

the three groups was as follows: 126 ng/g lwt (dismantling workers), 35.1 ng/g lwt (residents 

living within 50 km of the dismantling area), and 9.4 ng/g lwt (reference group). BDE 209 was 

the highest among all groups, explaining between 50 and 70% of total concentration. 

Interestingly, BDE 207 was the second highest in the electronic recycling group and the residents 

living within 50 km, explaining 8% and 15% of the total, respectively. Exceedingly high 

concentrations were found in one 18-year old male electronic waste worker: BDE 28, BDE 183, 

BDE 208, and BDE 209 were found at 148.3, 60.2, 66.2, and 3,436 ng/g lwt, respectively.

Bi et al. (2007) similarly found elevated levels of PBDEs in residents living near and 

working within an e-waste recycling facility in South China. They sampled two regions, 

including this recycling region and a second nearby region, where the fishing industry dominates 

the livelihood of the region. A total of 21 individuals were sampled within the recycling region 

and 26 from the nearby fishing region. Sixteen BDE congeners were measured, including 

BDE 209. Total concentrations from the recycling region ranged from 140 - 8,500 ng/g lwt, with 

a median of 600 ng/g lwt, while it more narrowly ranged from 16 - 490 ng/g lwt, with a median 

of 170 ng/g lwt, in the fishing region. BDE 209 dominated the concentrations, with a median of 

310 and maximum of 3,100 ng/g lwt in the recycling region and a median of 86 and maximum of 

370 ng/g lwt in the fishing region. Octa (BDE 197 and BDE 203) and nona (BDE 206, 

BDE 207, and BDE 208) congeners averaged from 20 to over 100 ng/g lwt in the serum of the 

residents from the e-waste recycling region, suggesting possible debromination of BDE 209. 

 Lee et al. (2007) measured 13 congeners (not including BDE 209) in 92 blood samples, 

including 30 from incinerator workers, 51 from nearby residents, and 11 from controls in 2001 

and 2002 in Korea. The average total concentration was 16.84 ng/g lwt, and there was only a 

slight difference between the incinerator workers, who had the highest concentrations at an 
5-86

average of 19.24 ng/g lwt, and the other two groups: residents at 15.22 ng/g lwt and controls at 

17.74 ng/g lwt. The difference between the groups was not significant, and there were no other 

correlations found, including to age, weight, dietary habits, and others, with the exception of sex: 

males had 15% higher levels compared to women. BDE 47 dominated the profile, explaining 

33% of the profile, followed by BDE 153 (24%), BDE 183 (17%), BDE 99 (15%), and 

BDE 100 (7%). 

5.6.6. Elevated Exposures at the High End of the General Adult Population

Table 5-7 shows the analysis of the 2003/2004 NHANES blood concentration data, 

which was described by Sjodin et al. (2008), used as a primary comparison to the predicted 

concentration in the model, and did not include statistics on total concentration, just statistics on 

individual congeners. Table 5-7 shows the congener-specific median concentrations, and, when 

added together, they lead to a total concentration of about 36 ng/g lwt. The statistics provided 

for individual congeners show that the 90th percentile concentration of all the key congeners in 

U.S. citizens are about 4−6 times higher than the 50th percentile and the 95th percentile is near or 

more than 10 times higher than the medians of the congeners. Perhaps more noteworthy was the 

finding that the highest total concentration found in an individual in NHANES 2003/2004 was 

3,680 ng/g lipid, which included BDE 47 at 2,350 ng/g lipid (Sjodin et al., 2008). This highest 

individual had concentrations about 100 times higher than the median in the population. Clearly, 

there are individuals with much higher concentrations than the central tendency median selected 

for the point estimate exercise of this chapter. 

 In comparison, this is not the same trend as generally found for dioxin. Ferriby et al. 

(2007) statistically evaluated the concentrations of polychlorinated dibenzo-p-dioxin (PCDD) 

and dibenzofuran (PCDF; the combination abbreviated CDD/F) concentrations from NHANES 

2001/2002. Determining the TEQ concentrations of the 17 toxic CDD/F congeners for 

individuals in the survey, they provided population statistics for CDD/F TEQ concentrations. 

The median concentration in the population was 14.4 pg/g lwt, the 95th percentile was 

45.2 ng/g lwt, and the maximum found was 139.2 ng/g lwt. In this case, the maximum found 

was 10 and not 100 times the median. To illustrate the difference in these NHANES results, 

Figure 5-3 shows the comparison of NHANES percentiles for CDD/F TEQ and BDE 47 findings 

at the 25th percentile, 50th percentile, 75th percentile, and 95th percentile, compared to the 
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Figure 5-3. The fraction of the maximum concentrations of PCDD/F TEQ 

and BDE 47 concentrations found at various percentiles within NHANES 

surveys of these two contaminants in adults. 

Note: NHANES data on PBDEs from Sjodin et al. (2008), and data on PCDD/F from Ferriby et al. 

(2007). 

maximum found in the NHANES surveys (dioxin data from Ferriby et al., 2007 and PBDE data 

from Sjodin et al., 2008). While the percentiles of dioxin concentrations within the general adult 

population, expressed as a fraction of the maximum found, range from between about 0.10 at the 

50th percentile and 0.30 at the 95th percentile for PCDD/Fs; they range much more narrowly at 

<0.01 at the 50th percentile to 0.07 at the 95% for BDE 47. 

Other studies show a similar disparity between the median (or geometric mean, if that is 

what the study authors provided) and high value of their body burden measurement study. For 

example, Morland et al. (2005) measured the blood of 93 individuals, 79 of which were anglers. 

The highest congener found was BDE 47, at a geometric mean of 13.3 ng/g lwt, followed by 

0

0.1

0.2

0.3

0.4

0.5

Fraction of Maximum Concentration

NHANES Percentiles Within Adult Population

25% 50% 75% 95% 100%

(1.00 )

PCDD/F TEQ

BDE 47

0

0.1

0.2

0.3

0.4

0.5

Fraction of Maximum Concentration

NHANES Percentiles Within Adult Population

25% 50% 75% 95% 100%

(1.00 )

PCDD/F TEQ

BDE 47
5-88

BDE 99, at 3.2 ng/g lwt. Although they did not report total concentrations, BDE 47 was found at 

a high of 1,388 ng/g lwt, which was 100 times the geometric mean for this congener, and the 

high BDE 99 concentration was 546 ng/g lwt, about 20 times the geometric mean. Anderson et 

al. (2008) provides summary information for blood sampling of 350 male and 158 female 

participants in a study of the relationships between several organic contaminants, including 

PBDEs, and consumption of Great Lakes sportfish. The geometric mean total concentration 

(sum of 8 congeners not including BDE 209) for males was 27 ng/g lwt and for females was 

25 ng/g lwt. They also showed a figure of the highest 50 concentrations for males and females. 

For males, the top five ranged from about 380 to 1,300 ng/g lwt (the top two at 1,100 and 

1,300 ng/g lwt) and for females, the top five ranged from about 250 to 950 ng/g lwt. Petreas et 

al. (2003) measured BDE 47 in adipose tissue from 32 women in the San Francisco area and 

found a median of 16.5 ng/g lwt, with a high of 510 ng/g lwt, 30 times the median. 

Evidence suggests that these elevated exposures at the very high end of the population 

could very well be due to dust exposures. Nearly every study that has been conducted measuring 

dust concentrations from different locations very similarly finds a reasonably log normal range 

until the very last few samples, which are substantially higher than the rest of the population. 

The following is a summary showing this trend from several dust studies: 

a. Stapleton et al. (2008b) measured BDEs on hands using hand-wipes. Their results and 

exposure estimates were described earlier in Section 5.3. The key finding for this 

discussion was that they measured BDEs in 33 individuals, and that for 32 of them, the 

measurement was under 500 ng/hand, but the 33rd individual had about 2,000 ng on his 

hand. The median from this population was 128 ng/hand. 

b. Allen et al. (2008a) collected 108 bulk dust samples from about 20 homes in the Boston, 

MA area, during two sampling events in 2006. Samples were characterized as having 

come from the “living room,” “bedroom,” and “vacuum bag” (meaning location 

unspecified). The geometric mean total concentrations found in the living room, 

bedroom, and vacuum bag were 13,732, 6,255, and 4,269 ng/g dry weight (dwt), 

respectively. The two highest total concentrations were 544,000 and 269,000 ng/g dwt, 

found in vacuum bag samples from a single individual’s home. The first high 

measurement was dominated by BDE 209, explaining 97% of the 544,000 ng/g dwt 

found; the dominant congener was not identified for the second highest sample. 

c. Two other studies of house dust in the United States showed similar trends of having a 

relatively consistent finding for most of the homes, and then one or two having 

substantially higher concentrations. In one, Stapleton et al. (2005) analyzed dust samples 
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from 17 homes in the Washington, DC area. They found total concentrations (including 

BDE 209) to be less than or near 7,000 ng/g dwt in 15 homes, but, in 2 homes, they found 

14,990 and 30,100 ng/g dwt, respectively. These samples were dominated by BDE 47, 

BDE 99, and BDE 209 in comparable amounts. In the other, Sharp and Lunder (2004) 

analyzed dust samples in 10 homes from nine states. Concentrations found in eight of the 

homes were near or less than 6,000 ng/g dwt total, but in two homes, the total 

concentrations were 16,000 and 41,000 ng/g dwt. These high samples were dominated 

by BDE 47, BDE 99, and BDE 209. 

d. Harrad et al. (2008a) measured 78 samples in four cities in Canada, New Zealand, the 

United Kingdom, and the United States. The median total concentration of 28 samples 

from the United Kingdom was 2,900 ng/g dwt, but 1 sample was found at 

520,000 ng/g dwt, with essentially all of it (>99%) explained by BDE 209. In another 

study, Harrad et al. (2008b) sampled dust in 30 homes, 18 offices, and 20 cars (68 total 

samples), and found 3 samples at 1,400,000, 2,200,000, and 2,600,000 ng/g dwt, 

essentially all of it being BDE 209. 

 Perhaps a certain behavior within a household or office that puts one in close contact with 

a product containing BDEs or in close contact with dust heavily laden with BDEs explains this 

tendency in populations to have very high exposures at the high end. This is clearly an 

uncertainty that requires further investigation. Besides dust, the cause for these higher exposures 

could also be a high consumption rate of a particular food product that may be impacted by local 

conditions. For example, Sjodin et al. (2000) found that Swedish fishermen consuming large 

amounts of Baltic Sea fish had a median BDE 47 level five times higher than that of 

nonconsumers. However, only a small statistically insignificant difference was found between 

urban anglers and nonanglers in a study of 93 individuals (79 anglers, 14 nonanglers) in the 

United States (Morland et al., 2005). It should also be noted that findings of PBDE body 

burdens in other U.S. studies are not very different from the NHANES study. As described in 

Section 5.2., the means or medians from essentially all U.S. studies are within a reasonably 

narrow range of PBDE concentrations: 30 to 100 ng/g lwt. The consistency in central tendency 

findings in these studies suggests consistency in exposure of Americans to PBDE. 

A simple exercise may be able to shed light on the cause of extremely elevated exposures 

to a small number of individuals at the high end of measured populations in the United States. 

The implication from discussions above, of course, is that dust-related exposures are the likely 

cause of such elevated blood levels. But can that be demonstrated, and can food exposures also 

possibly explain these high levels? Sjodin et al. (2008) note a 95th percentile blood concentration 
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of BDE 47 of 157 ng/g lwt in NHANES 2003/2004. With the use of eq 5-3 in Section 5.5 above, 

one can calculate the dose required under an assumption, for example, that the elevation was due 

entirely to dust ingestion or food ingestion. Starting with dust ingestion, the dose corresponding 

to this blood concentration, assuming k of 0.00063 day−1 (half-life of 3 years), a body lipid 

mixing reservoir of 17,500 g, and an absorption of 0.69 (used for BDE 47 absorbed to dust) 

would be 2,508 ng BDE 47/day, or for a 70-kg adult, 35 ng/kg-day. If it was from food 

ingestion, the absorption would be higher at 0.94, and the dose would be 1,841 ng BDE 47/day, 

or 26 ng/kg-day. The key uncertain parameter in this calculation is the elimination rate, which 

was based on a half-life of 3 years. Had the half-life been 6 years instead of 3 years, the doses 

would be halved at 18 ng/kg-day for dust and 13 ng/kg-day for food. These backcalculated 

doses can now be compared to forward calculated doses of BDE 47 from dust and from food. 

Allen et al. (2008a) found a maximum BDE 47 concentration in dust in Boston of 

16,840 ng/g dwt, Harrad et al. (2008a) found a high BDE 47 concentration in Texas dust of 

3,300 ng/g dwt, Sharp and Lunder (2004) found a high of 9,070 ng/g dwt in Montana dust, and 

Rudel et al. (2003) found a high of 9,860 ng/g dwt in Cape Cod dust. If one assumes an ongoing 

exposure to a “high” concentration of 10,000 ng/g dwt BDE 47 in dust and an average ingestion 

rate of 50 mg/day, then the dose on a body weight basis would be 7.1 ng/kg-day, and if one then 

calculates soil dermal contact in the manner done in this report, this would bring the total to 

8.9 ng/kg-day. This is below the backcalculated level for hypothetical exposure via dust of 35 

(or 18 at a 6-year half-life) ng/kg-day. But for food, the disparity may be larger. Schecter et al. 

(2006b) found high concentrations of BDE 47 of 387 pg/g wwt in meats, 2,748 pg/g wwt in fish, 

and 105 pg/g wwt in dairy. If these concentrations are used to calculate food exposures to 

BDE 47 using the contact rates in this assessment, than the total adult dose would be 115 ng 

BDE 47/day, or 1.6 ng/kg-day, well below the backcalculated food dose of 26 (or 13) ng/kg-day. 

Higher exposures would be calculated using higher contact rates, but also, the maximum BDE 47 

found in NHANES 2003/2004 is 2,350 ng/g lwt (Sjodin et al., 2008). It appears as if exposures

other than food and dust that are not identified or quantified properly in this assessment are 

affecting a small percentage of individuals at the high end (upper 5%, perhaps) of populations, at 

least for BDE 47. 
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5.7. UNCERTAINTY AND VARIABILITY IN ESTIMATING INTAKE DOSE AND 

CONVERTING THAT DOSE TO A BODY BURDEN

The analysis in this chapter has taken a “point estimate” approach in the intake 

estimation, the pharmacokinetic modeling, and the comparison with measured body burdens. In 

so doing, it has arrived at a fairly broad-reaching finding that the bulk of exposures are 

indoor-dust related. The exercise revealed that about 7 ng/kg-day of total PBDE exposure 

appears necessary in order to reproduce the median body burdens seen in the adult population. 

This finding is based on pharmacokinetic modeling. Given the model and parameters chosen for 

BDE congeners, the median body burden would not be duplicated unless the intake dose were 

near this value of 7 ng/kg-day. The point estimate approach suggested that about 90% of this 

total intake came from dust ingestion, inhalation, and dermal contact with dust. However, there 

are uncertainties throughout this exercise, as well as variabilities in exposure. As discussed in 

the previous section, there appears to be exposures at the very high end of the general population 

that are substantially higher than the central tendency exposures that are characterized by a dose 

in the range of 7 ng/kg-day, dominated by dust exposures. Further, it appears that these very 

high body burdens cannot be explained even by exposures to the highest dust concentrations and 

food concentrations found. 

The point estimate construct of the exercise does not guarantee that the finding of the 

importance of the dust pathways to mid-range (such as median) body burdens is proven. The 

purpose of this section is to identify and discuss known uncertainties and variabilities so that this 

primary finding and other findings in this chapter can be understood in their proper context. 

5.7.1. Uncertainties with Estimates of Dust Intakes of PBDE

 The amount of 50 mg/day assumed for adult soil/dust ingestion is a classic uncertainty. 

Described as an “average” value for soil ingestion in the Exposure Factors Handbook

(U.S. EPA, 1997), the actual amount of ingestion of house dust plus outdoor soil could very 

easily be much lower. While Sjodin et al. (2004b) assumed a dust ingestion rate of 100 mg/day, 

Stapleton et al. (2005), on the other hand, assumed an adult house dust ingestion rate of 0.56 

mg/day, over 2 orders of magnitude lower, explaining their estimate of 3.3 ng/day of exposure to 

PBDEs via ingestion exposure to household dust. As noted earlier in the chapter, this value of 

0.56 mg/day house dust was listed in EPA’s Exposure Factors Handbook (U.S. EPA, 1997), 
5-92

which cited Hawley (1985) for using a value of 0.56 mg/day to characterize adult exposure to 

house dust from normal activities in the house (higher exposures of over 100 mg/day resulted 

from “work in the attic”). As noted, Stapleton et al. (2005) estimated an exposure to total 

PBDEs of 3.3 ng/day based on this assumption. It would seem that their subsequent studies, and 

those of others, essentially prove that this is much too low an estimate for house dust ingestion, 

at least in the context of estimating exposure to house dust. Specifically, their recent study 

(Stapleton et al., 2008b) directly measuring BDEs on hands using wipes, and then using an 

empirical model to estimate hand-to-mouth exposures, found a median exposure estimate for 

adults of 154 ng/day. The soil/dust ingestion rate of 100 mg/day used in the characterization of 

children’s exposure (ages 1−5) is similarly uncertain. In the recent Child-Specific Exposure 

Factors Handbook (U.S. EPA, 2008), however, the recommended factors and approach for 

modeling childhood (ages 1−6) exposures to outdoor soil and indoor house dust is really not that 

much different than used here. Specifically, the current approach, as noted, uses a total soil/dust 

ingestion rate of 100 mg/day, where the concentration on that matrix is a weighted average of 

outdoor soil (about 20%) and indoor house dust (about 80%). Similarly, the Child-Specific 

Exposure Factors Handbook recommends a central tendency 60 mg/day of house dust ingestion 

using, presumably, concentrations on house dust, and 30 mg/day of outdoor soil using, 

presumably, concentrations in outdoor soil. 

The procedure here to characterize ingestion of house dust is to multiply a total amount 

ingested by a fraction that comes from the house. The assumption of 0.90 for this fraction, based 

on data suggesting 90% of the time is spent indoors, is conservative. Paustenbach et al. (1997) 

looked at data suggesting that 50% of house dust originates from outdoor soil. This may be less 

important for the current exercise because dust concentrations were taken from house dust 

measurements, so the origin of the dust is not relevant when using measured dust concentrations 

directly. Still, the assumption that 90% of the ingested amount is house dust is really not 

substantiated in the literature. It is appropriate for “time spent indoors” and, for obvious reasons, 

would be reasonable for inhalation exposures. 

Bioavailability was considered in the context of absorption fractions for dust ingestion 

and dust dermal contact. In the dermal contact pathway, an absorption fraction of 0.03 (3% 

absorbed) was assumed for all congeners (see Table 5-5), and in the dust ingestion pathway, 

different absorption fractions for each congener ranging from about 0.20 to 0.80 were assumed. 
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There is a subtle difference in the way in which results are presented in this study, in that for all 

pathways except dermal contact, the intake dose is calculated just by the contact rate (food 

ingestion rate, inhalation rate, etc.) multiplied by the concentration of BDEs in the contact media 

(in food, air, etc.). For the dermal contact pathway, this intake dose has already considered this 

absorption fraction of 0.03. For the other pathways, the absorption fractions are used in the PK 

modeling of body burdens due to intake doses; the absorption fractions reduce the intake dose to 

consider absorbed dose. In any case, all of the absorption fractions are uncertain. The value of 

0.03 for dermal contact was used in the modeling of background exposures to dioxin based on 

literature showing that this tightly-sorbed contaminant would not desorb readily from soil 

contacting the skin and then penetrate the skin surface to a great extent (U.S. EPA, 2003). 

However, this may not be true for PBDEs, or at least true for the case where the vehicle is house 

dust rather than soil. Only a few studies were found that have measured dermal absorption of 

PBDEs. Although none of these used PBDEs sorbed to dust, soil, or other solid matrices, they 

provide some indication of the dermal absorption potential of these chemicals. Hughes et al. 

(2001) examined the in vitro dermal absorption of [14C] decabromodiphenyl oxide. Skin from 

the adult hairless female mouse was removed and mounted in flow-through diffusion cells. The 

chemical was applied to the skin at three dose levels (6, 30, and 60 nmol) in a volatile vehicle 

(tetrahydrofuran). The 24-hour cumulative percent of the dose in the receptor fluid was 

0.07−0.34%. The percent of the applied dose detected in the skin after 24 hours ranged from 2 to 

20%. Staskal et al. (2005) conducted a mouse in vivo study to measure dermal absorption of 

2,2’,4,4’-tetrabromodiphenyl ether (BDE 47). A single dose was applied to the skin in an 

acetone solution. About 62% of the administered dose was absorbed over a 5-day period. Roper 

et al. (2006) studied the dermal absorption of 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47) using 

rat and human skin in vitro. A single dose was applied to the skin in an acetone solution. The 

total absorbed dose in human skin after 24 hours was 3.13% (1.93% in the receptor solution and 

1.2% in the skin). The total absorbed dose in rat skin after 24 hours was 17.94% (14.81% in the 

receptor solution and 3.13% in the skin). Roper et al. (2006) showed that rat skin was more 

permeable than human skin to TBDE, and this has also been observed for other chemicals and 

other rodents in multiple studies (i.e., van Ravenzwaay and Leibold., 2004). All three studies 

applied the chemical in a solvent that evaporated rapidly, leaving a residue of pure chemical on 

the skin. This would tend to increase the absorption relative to a similar dose that has been 
5-94

absorbed to dust. Thus, the 3% absorption based on human skin in vitro testing by Roper et al. 

(2006) is probably the most relevant datum, but may be high for dust exposures. U.S. EPA 

(2004) recommends 3% absorption for TCDD in soil. The similarity of TCDD to PBDEs in 

terms of lipophilicity and molecular size adds support to the assignment of 0.03 for PBDEs. The 

absorption fractions for dust ingestion were derived from a study by Huwe et al. (2008), where 

BDE congeners were administered to male rats in corn oil and household dust. It was found that 

absorption amounts were similar for corn oil and dust, and the congener-specific results from 

that study, showing absorption fractions ranging between 0.18 and 0.78, were directly used for 

the dust ingestion pathway in this study. The absorption fractions for food ingestion and 

inhalation were higher, at between 0.78 and 0.94. Organic compounds sorbed to soil or dust are 

less bioavailable than when ingested in food or inhaled. Paustenbach et al. (2006) examined the 

literature on the bioavailability of 2,3,7,8-TCDD. In their Monte Carlo simulations on 

contaminated soil exposures, they assumed that the oral absorption of this compound in 

contaminated soil followed a lognormal distribution with a range of 0.5 to 63% and a mean value 

of 35%. ATSDR (2004) reviewed the literature on bioavailability of PCBs in soil and concluded 

that a range of 40−65% was appropriate. Paustenbach et al. (1997) did comment on the fact that 

the bioavailability of contaminants on house dust, in general, is much greater than that in outdoor 

soil, because house dust particles are finer, therefore containing more surface area, than soil 

particles. In a backward pharmacokinetic modeling exercise in which McDonald (2005) derived 

intake estimates that would correspond to measured body burdens, he assumed absorption 

fractions of 0.78 to 0.94 for BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154. This was not 

pathway-specific and was based on absorption experiments in rats, and although the carrier in the 

experiments was not noted, it was unlikely to be soil. EPA assumed that the absorption fraction 

of dioxin from a total intake dose was 0.80 (U.S. EPA, 2003), but dioxin intakes are dominated 

by food―not dust/soil. In general, the point is made that while an intake dose of PBDEs could 

be dominated by dust ingestion and dermal contact with dust, there remains uncertainty as to 

how much of the doses by these pathways get absorbed to eventually appear in blood. 

5.7.2. Uncertainties and Variabilities with Other Pathways

 There is less uncertainty in characterizing central tendency for water ingestion, food 

ingestion, and inhalation, as compared to soil/dust ingestion and soil/dust dermal contact. Food 
5-95

intake quantities as developed in the Exposure Factors Handbook (U.S. EPA, 1997) were 

derived from the U.S. Department of Agriculture Continuing Survey of Intakes by Individuals, 

which is a survey of high quality used in many assessments, such as the U.S. Food and Drug 

Administration’s market basket surveys, which are used to determine intakes from measurements 

of contaminants in sampled food. Water ingestion and air inhalation are also comprehensively 

studied, and the contact rates chosen are characterized as very reasonable central tendency point 

estimates. Furthermore, most food surveys arrive at comparable concentrations in food products. 

One exception described in Chapter 4 was the finding by the UK’s Food Standards Agency of 

high BDE 209 in all food sampled (FSA, 2006). The quality of these data are unknown and 

might be questioned since it differs from essentially all other studies of PBDEs in food. When 

combining food concentrations of PBDEs with food intakes, most studies around the world, 

including the United States, have arrived at adult food intakes in the range of 1−2 ng total 

BDE/kg-day. Inhalation and water ingestion have mostly arrived at adult exposures of less than 

1 ng/kg-day. These findings can easily be seen in Table 5-4, which lists intake estimates from all 

pathways from studies around the world. Only the UK FSA study showing unusually high 

BDE 209 concentrations arrived at food intakes greater than this 1−2 ng total BDE/kg-day range. 

 It should also be noted that there may be pathways not considered in this assessment. For 

example, while the assessment considered dermal contact with dust, it did not consider dermal 

contact with PBDE-treated products. Dermal absorption of PBDEs may occur via direct contact 

with treated materials such as clothing, carpeting, upholstery, etc. Wester et al. (1996) has 

shown that chemicals in fabric can transfer from fabric into and through human skin. This was 

based on human skin in vitro experiments with glyphosphate and malathion applied to cotton 

sheets. Absorption from dry cloth was found to occur, but it was less than the chemicals in 

aqueous solutions. When the cloth was wetted with water to simulate sweating, absorption 

increased. For example, absorption of malathion from aqueous ethanol solution was 8.77 

±1.43%. This decreased to about 0.60% for dry cotton sheets. However, absorption from cotton 

sheets increased to 7.34 ±0.61% when wetted with aqueous ethanol. The discussions above on 

the NAS and VCCEP submissions did, in fact, discuss mouthing of products including treated 

fabrics, and, also, the direct contact with treated fabrics. The exercises were specific to 

decaBDE, were conservative by design, but showed much higher exposures than the central 

tendency estimates presented here (see Section 5.6.4 for more detail). Finally, Stapleton et al. 
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(2008b) provides estimates for a direct hand-to-mouth pathway. They estimated the amount of 

PBDEs on the hands of 33 volunteers using hand wipes, and using hand-to-mouth, estimated 

intakes for children ranged as high as 6,090 (95%) ng total BDE/day, with a median of 1,390 ng 

total BDE/day. 

 The possibility of exposure by direct contact with treated surfaces, or surfaces that could 

be reservoirs of PBDEs otherwise released from their original sources, is being studied by a 

novel approach using X-ray fluorescence (XRF). XRF analyzers measure the bromine 

concentrations on the dust that resides on surfaces of household items. Allen et al. (2008b) 

validated their use on household items by showing that XRF readings were highly correlated 

with GC/MS measured bromine and PBDE levels in furnishings and electronic devices 

(R = 0.93). Imm et al. (2009) took this application a step further by measuring bromine 

concentrations on the surfaces of many household items for a cohort in Wisconsin and then 

further measuring the PBDEs in blood of residents who lived in the homes they studied. They 

found that the bromine content in the participant’s sleeping pillow (p-value = 0.005) and primary 

vehicle seat cushion (p-value = 0.03) were the strongest predictors of PBDE blood 

concentrations.

 

5.7.3. Uncertainties with Pharmacokinetic Modeling

The choice of the simple one-compartment 1st order model is reasonable for PBDEs. 

Like dioxins, for which the model has been extensively and successfully used (U.S. EPA, 2003, 

and other citations not provided), most PBDEs are lipophilic and persistent in the body (there is 

some evidence and discussion in this report that BDE 209 is not as lipophilic as other BDE 

congeners, and its elimination half-life is hypothesized to be on the order of days, in contrast to 

years for lower-brominated BDEs). The application of the model at steady state instead of in a 

temporally variable mode could introduce uncertainties. For dioxins, it was found that 

application of the model at steady state using dose estimates developed for current conditions 

would underestimate average adult body burdens by about one-half (U.S. EPA, 2003). This is

because of high dioxin intakes in the middle decades of the twentieth century (the 1960s until 

about 1980; much higher than current intakes), such that body burdens of older adults are higher 

than younger adults, driving up the current average adult population body burden. Similar age 

trends were not found in BDE population studies. Rather, the younger age ranges were found to 
5-97

have higher body burdens in some studies. Unlike dioxins, PBDEs were not in the environment 

prior to their introduction into household products beginning in the 1970s, and body burdens 

were first found to contain PBDEs in the 1980s. With the voluntary withdrawal of pentaBDE 

and octaBDE formulations in 2004, there may be declines seen in the key congeners of these 

formulations in the most recent surveys. Generally, though, it may be reasonable to assume that 

exposures have at least remained steady if not increased (perhaps because of rising levels in dust 

as more products using the PBDEs were incorporated into modern living) between 1980 and the 

present. This provides a reasonable justification for use of the simple PK model in a steady-state 

mode, at least in comparison with dioxin, where steadily declining exposures leads to an 

underestimate of population body burdens if using a steady-state model 

The uncertainties associated with the dose estimates were discussed in the previous 

section. Absorption was also discussed in the previous section in the context of absorption of 

PBDEs from ingested dust. There was limited choice in the literature from which to make 

selections of elimination half-lives. Only one study, Geyer et al. (2004) was found to assign 

half-lives to BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154 of 2.9 to 11.0 year. In 

addition, only one study was found, Thuresson et al. (2006), which provided estimates for 

BDE 183 of 0.26 year (94 days) and for BDE 209 of 0.041 year (15 days). 

With these selections, it was found that BDE 47 appears to be underpredicted. The 

prediction was 8.6 ng/g lwt, compared to observations of 20.5 ng/g lwt in blood and 26.0 ng/g 

lwt in breast milk. The cause for this underprediction is not known, but it could very easily be 

the assumed half-life in humans. At 3.0 years, it dissipated nearly twice as fast as BDE 99, 

which was assigned a half-life of 5.4 years. Had it dissipated at a half-life of 10 years, the 

prediction would jump to greater than 30 ng/g lwt, now more than twice the prediction for

BDE 99, and more in line with current measurements. In the same vein, BDE 99 was 

overpredicted. It was predicted at 12.5 ng/g lwt, while is measured as 5.0 ng/g lwt in blood and 

5.4 ng/g lwt in breast milk. Perhaps the half-lives of both congeners should be reversed. 

However, other evidence in the literature supports the assumption that BDE 47 is eliminated 

more rapidly than BDE 99. In an experiment where BDE 47, BDE 99, BDE 100, and BDE 153 

were all administered intravenously to mice, Staskal et al. (2006) found that tissue concentrations 

were highest for BDE 153, followed by BDE 100, BDE 99, and BDE 47. Similar to the human 

data, these mouse data suggest a more rapid elimination of BDE 47 as compared to BDE 99 and 
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also other key congeners. On the other hand, the human body burden data show that BDE 47 

concentrations are the highest of all congeners and about four times higher than BDE 99 

concentrations. The modeling in this study (given uncertainties of course) suggests that the dose 

of the two congeners is about equal. Logically, therefore, one would speculate that BDE 99 

would be more rapidly eliminated as compared to BDE 47, but that is not what the human 

toxicokinetic data, or the rodent data cited, have found. This trend of higher BDE 47 

concentrations in humans may suggest debromination of BDE 99 (or higher-brominated 

congeners) to form BDE 47 in the body. In other words, the modeling may correctly determine

body burdens due to intakes, but the half-lives assigned might not reflect full elimination from 

the body but rather higher-brominated congeners are being transformed to lower-brominated 

congeners. However, this is much too speculative given current information on human and 

animal metabolism. It might be possible that the one study measuring the elimination of these 

congeners in an occupational cohort was not appropriate for the half-lives of these two congeners 

in the background population. It would be premature to use this framework to “calibrate” 

congener-specific half-lives in humans. Still, it gives an indication of where key information 

gaps are. 

With regard to the final total concentrations, they were predicted at 31.0 ng/g lwt and 

measured at 36.3 ng/g lwt in blood and 44.1 ng/g lwt in breast milk. Other studies in breast milk 

showed different concentrations, some lower and some higher. Most other blood measurement 

studies were somewhat higher than this NHANES result, but most studies found central tendency 

total concentrations to be well under 100 ng/g lwt, so the prediction is about within a factor of 2 

of most measurements, and in the case of NHANES blood and the selected breast milk 

concentrations, predictions came very close to measurements. 

The results of this simple PK modeling exercise are consistent with those of McDonald 

(2005), although he took the opposite approach: he started with body burdens and used reverse

PK modeling to derive the intake doses that would explain the body burdens. He used the same 

PK parameters (absorption fractions and congener-specific half-lives) as were used in this 

exercise, although he did not separately consider dust ingestion with a different absorption 

fraction. Compiling data on PBDEs in human milk (including some of the same data 

summarized above), he found a median concentration of 48 ng/g lwt, a mean of 90 ng/g lwt, and 

a 95% of 302 ng/g lwt. These totals were the sum of BDE 47, BDE 99, BDE 100, BDE 153, and 
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BDE 154. Modeling these congeners individually, he found that the doses of total PBDEs that 

would explain these body burdens were 8.5, 16.0, and 53.6 ng/kg-day for the median, mean, and 

95%. In the exercise above, dose was forward calculated from exposure media concentrations 

and contact rates, and the adult dose was estimated at 500 ng/day total, or on a body weight basis 

assuming a 70-kg adult, 7.1 ng/kg-day. This dose was estimated based on “representative” 

central tendency media concentrations and average contact rates, so the consistency of 

7.1 ng/kg-day derived by a forward calculation for “average” conditions with McDonald’s 

(2005) 8.5 ng/kg-day derived by a retrocalculation from the median body burden is very 

encouraging. 

 McDonald (2005) cites European body burdens, which are more like 10 ng/g lwt, as well 

as estimates of dose made by Europeans based on food alone, which are in the range of 

1.0 ng/kg-day, in an application of his PK framework. With his model, McDonald (2005) found 

quantitative consistency in those two quantities: i.e., the dose he back calculated from body 

burdens matches the forward calculated dose by food ingestion developed by European 

researchers. In essence, he is implying that the primary pathway of BDE exposure to Europeans 

is through food consumption. It is unclear why house dust and other indoor sources of PBDEs 

were not considered. These pathways were considered by Harrad et al. (2006) in a study 

conducted in the United Kingdom. They observe that dust concentrations are much lower in the 

United Kingdom as compared to the United States, while food concentrations are comparable in 

the two countries. They suggest that the higher body burdens found in the United States are 

likely due to house dust exposures. If not entirely due to house dust, there may be other 

pathways of exposure Americans have that Europeans do not. In any case, the prevalence of 

lower-brominated BDEs in dust in the United States and the subsequent impact on U.S. body 

burdens is a key issue identified in this assessment, and, needless to say, it requires further 

research.

5.8 OVERALL FINDINGS OF EXPOSURE OF AMERICANS TO PBDEs 

Examination of literature data and exposure exercises in this chapter supports these 

general findings: 
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1. PBDEs bioaccumulate in lipids: Most PBDEs bioaccumulate in lipids and typically 

body burden measurements are expressed in a ng/g lipid wt basis. The finding that 

PBDEs partition into lipids is upheld for congeners up to the hexa congeners as studies 

show similar concentrations measured in maternal blood and maternal milk for the 

commonly measured congeners, BDE 28, BDE 47, BDE 99, BDE 100, BDE 138, 

BDE 153, and BDE 154. There is some evidence that the hepta congener, BDE 183, and 

more so, the deca congener, BDE 209, are found at lower concentrations in mother’s milk 

than mother’s blood. Concentrations of BDE 209 found in mother’s milk in the United 

States are in the range of 0.5 ng/g lwt when quantified, while it is near 5 ng/g lwt in blood 

when quantified. 

2. PBDE Body burdens in U.S. adults: Total PBDE body burdens in Americans are in the 

range of 30 to 100 ng/g lwt. The most statistically rigorous and expansive study of 

general population exposures to PBDEs in the United States is an analysis of 2003/2004 

NHANES data by Sjodin et al. (2008). The geometric mean congener concentrations 

over the entire population above the age of 12 were, in ng/g lwt (descending order): 20.5 

for BDE 47, 5.7 for BDE 153, 5.0 for BDE 99, 3.9 for BDE 100, and 1.2 for BDE 28. 

The sum of these geometric means was 36.3 ng/g lwt. The 95th percentile for the total 

body burden of PBDEs (i.e., sum of these five congeners) was 291 ng/g lwt, and the 

maximum found was 3,680 ng/g lwt. The predominant congener found in nearly all body 

burden studies in the United States is BDE 47, explaining about 50% of the total 

concentration. The second most commonly found congeners are BDE 99 and BDE 153, 

both explaining in the range of 10−20% of total concentrations. Most of the studies have 

not measured BDE 209, but when measured, it was generally found in about half the 

samples at low levels near 5 ng/g lwt.

3. PBDE Body Burdens in other Countries: Body burdens of Americans are higher than 

body burdens of individuals in other countries. Body burdens in other countries are less 

than 10 ng/g lwt. Most non-U.S. data are from Europe. 

4. PBDE Body Burdens in U.S. Children: Body burden data, as well as intake and body 

burden impact modeling, suggest that the infant and toddler have higher exposures than 

older children or adults. In paired sampling studies, including mother and child, the 

child’s body burden has exceeded by the mother’s by about a factor of 4, exceeding 100 

ng/g lwt total PBDEs in some cases. Higher levels in toddlers were attributed to 

ingestion of mother’s milk and high exposures to house dust. 

5. PBDE Fetal Body Burdens: Limited data from fetal tissue and numerous studies 

including measurements of BDE congeners from umbilical cord blood support the 

conclusion that the fetus is exposed to PBDEs through the mother. In some paired 

studies including both maternal and umbilical cord serum, there is a suggestion that the 

fetus may be less exposed than the mother as lower concentrations of BDE congeners in 

umbilical cord as compared to maternal serum were found. This trend of lower 

concentrations were seen mostly for higher-brominated congeners in two of these studies, 
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where this general observation could be made, but in four other studies reviewed, similar, 

if not slightly higher, concentrations of all congeners were seen in umbilical cord serum. 

6. High-End Exposures in U.S. Population: An important trend that warrants further 

investigation is that, even in background adult populations, there are individuals 

experiencing very high exposures. This has been seen in studies of PBDEs in blood as 

well as house dust measurement studies, suggesting that dust exposures could explain 

these unusually high exposures. However, modeling exercises undertaken to examine 

this hypothesis suggest that even the highest dust concentrations might not be able to 

explain the highest body burdens found. As such, these very high body burdens remain a 

critical uncertainty that warrants further investigation. 

7. Occupational Exposure: Limited occupational data support the observation that 

individuals in occupations that would lead to higher exposures to specific congeners have 

higher concentrations of PBDE congeners in their blood than the general population. 

8. PBDE Intake Rates: Intakes have been expressed as a straight-dose basis, ng total 

PBDE/day, or on a body-weight basis, ng total PBDE/kg-day. Intake estimates in the 

literature have tended to focus more on intake by food than by house dust, although this 

has changed in recent years as researchers recognize the importance of house dust in the 

overall exposure paradigm for PBDEs. Total daily intake estimates derived in this study, 

based on exposure media concentrations derived in Chapter 4 and combined with average 

contact rates, were 500 ng/day for adults to 1,411 ng/day for nursing infants. Assuming 

an average weight of 70 kg, the estimated adult intake rate was 7.1 ng/kg-day. These 

intakes were driven by indoor exposures via soil/dust ingestion, dermal contact with dust, 

and inhalation of indoor air; those three pathways accounted for about 90% of total 

intakes, with food and water ingestion explaining the remaining 10%. 

9. Intake Rates for Infants and Children: Intake modeling for the breast milk pathway 

combining measured milk concentrations and infant ingestion of human milk led to an 

intake of 1,411 ng/day in this study. Assuming an average body weight of 10 kg for an 

infant during the months of breast-feeding, a dose is calculated as 141 ng/kg-day. It is 

lower than intake estimates for children derived in this study, which were 47.2 ng/kg-day

for ages 1−5 (assuming 15 kg bw), 13.0 ng/kg-day for ages 6−11 (assuming 30 kg bw), 

and 8.3 ng/kg-day for ages 12−19 (assuming 58 kg bw). 

10. PBDE Dust Intake rates: Estimates in the open literature for intake of total PBDEs 

(sum of congeners, with most estimates for dust including BDE 209) via house dust 

ingestion range from 3 ng/day to 400 ng/day. The latter assumed less than 1 mg/day dust 

ingestion, while the former assumed a 100 mg/day dust ingestion rate. The adult dust 

intake rate derived in this study (dust ingestion plus dermal absorption) was 449 ng/day. 

 

11. PBDE Dietary Intake Rates: Estimates in the open literature of total PBDE intakes 

from food ingestion were in the range of 0.5 to 2.0 ng/kg-day for adults. 
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12. Modeling PBDE Body Burdens from Intake Rates: Using a simple PK model 

parameterized with available literature values, adult lipid-based concentrations (not 

specific to blood or milk) were predicted, starting with the intake values derived for 

adults in this study (see bullet 8 above). On a total PBDE basis, the prediction was low at 

31.0 ng/g lwt, while it was observed at 36.3 ng/g lwt in blood and 44.1 ng/g lwt in milk 

as central tendency values in studies selected as representative of the general population. 

Specifically, congener-specific geometric means from NHANES 2003/2004 in blood 

were summed to get the total concentration, and congener-specific medians from 

opportunistic sampling in breast milk were summed to get the total concentration. 

Predictions were reasonably close to measurements for five of eight congeners. While 

these predictions encouragingly match observations for the majority of congeners, 

uncertainties exist in the exercise, starting from development of dose estimates based on 

limited environmental measurements, to indoor contact rates with house dust, to the PK 

parameters of absorption and elimination half-life. There is also variability in U.S. body 

burdens. 

13. The Importance of the Dust Pathway: The overall weight-of-evidence of this exercise 

supports the finding that the bulk of U.S. exposures occur in the indoor environment 

through contact with house dust. The exercise suggests contact with house dust accounts 

for between 80 and 90% of total exposures, with the remainder due primarily to food 

ingestion. Circumstantial evidence supporting this hypothesis was the high 

concentrations found in U.S. house dust, and other researchers have also identified house 

dust as a key matrix of exposure concern for these compounds. 

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