TOXICOLOGICAL PROFILE FOR 

POLYBROMINATED BIPHENYLS AND 

POLYBROMINATED DIPHENYL ETHERS 

U.S. DEPARTMENT OF HEALTH AND HUMAN SERVICES 

Public Health Service 

Agency for Toxic Substances and Disease Registry 

September 2004 
PBBs and PBDEs ii 

DISCLAIMER 

The use of company or product name(s) is for identification only and does not imply endorsement by the 

Agency for Toxic Substances and Disease Registry. 
PBBs and PBDEs iii 

UPDATE STATEMENT 

A Toxicological Profile for PBBs and PBDEs, Draft for Public Comment, was released in 2002. This 

edition supersedes any previously released draft or final profile. 

Toxicological profiles are revised and republished as necessary. For information regarding the update 

status of previously released profiles, contact ATSDR at: 

Agency for Toxic Substances and Disease Registry 

Division of Toxicology/Toxicology Information Branch 

1600 Clifton Road NE, 

Mailstop F-32 

Atlanta, Georgia 30333 



PBBs and PBDEs vii 

QUICK REFERENCE FOR HEALTH CARE PROVIDERS 

Toxicological Profiles are a unique compilation of toxicological information on a given hazardous 

substance. Each profile reflects a comprehensive and extensive evaluation, summary, and interpretation 

of available toxicologic and epidemiologic information on a substance. Health care providers treating 

patients potentially exposed to hazardous substances will find the following information helpful for fast 

answers to often-asked questions. 

Primary Chapters/Sections of Interest 

Chapter 1: Public Health Statement: The Public Health Statement can be a useful tool for educating 

patients about possible exposure to a hazardous substance. It explains a substance’s relevant 

toxicologic properties in a nontechnical, question-and-answer format, and it includes a review of 

the general health effects observed following exposure. 

Chapter 2: Relevance to Public Health: The Relevance to Public Health Section evaluates, interprets, 

and assesses the significance of toxicity data to human health. 

Chapter 3: Health Effects: Specific health effects of a given hazardous compound are reported by type 

of health effect (death, systemic, immunologic, reproductive), by route of exposure, and by length 

of exposure (acute, intermediate, and chronic). In addition, both human and animal studies are 

reported in this section. 

 NOTE: Not all health effects reported in this section are necessarily observed in the clinical 

setting. Please refer to the Public Health Statement to identify general health effects observed 

following exposure. 

Pediatrics: Four new sections have been added to each Toxicological Profile to address child health 

issues: 

 Section 1.6 How Can (Chemical X) Affect Children? 

 Section 1.7 How Can Families Reduce the Risk of Exposure to (Chemical X)? 

 Section 3.7 Children’s Susceptibility 

 Section 6.6 Exposures of Children 

Other Sections of Interest: 

Section 3.8 Biomarkers of Exposure and Effect 

Section 3.11 Methods for Reducing Toxic Effects 

ATSDR Information Center 

 Phone: 1-888-42-ATSDR or (404) 498-0110 Fax: (770) 488-4178 

 E-mail: atsdric@cdc.gov Internet: http://www.atsdr.cdc.gov 

The following additional material can be ordered through the ATSDR Information Center: 

Case Studies in Environmental Medicine: Taking an Exposure History—The importance of taking an 

exposure history and how to conduct one are described, and an example of a thorough exposure 

history is provided. Other case studies of interest include Reproductive and Developmental 

Hazards; Skin Lesions and Environmental Exposures; Cholinesterase-Inhibiting Pesticide 

Toxicity; and numerous chemical-specific case studies. 
PBBs and PBDEs viii 

Managing Hazardous Materials Incidents is a three-volume set of recommendations for on-scene 

(prehospital) and hospital medical management of patients exposed during a hazardous materials 

incident. Volumes I and II are planning guides to assist first responders and hospital emergency 

department personnel in planning for incidents that involve hazardous materials. Volume III—

Medical Management Guidelines for Acute Chemical Exposures—is a guide for health care 

professionals treating patients exposed to hazardous materials. 

Fact Sheets (ToxFAQs) provide answers to frequently asked questions about toxic substances.

Other Agencies and Organizations 

The National Center for Environmental Health (NCEH) focuses on preventing or controlling disease, 

injury, and disability related to the interactions between people and their environment outside the 

workplace. Contact: NCEH, Mailstop F-29, 4770 Buford Highway, NE, Atlanta, 

GA 30341-3724 • Phone: 770-488-7000 • FAX: 770-488-7015. 

The National Institute for Occupational Safety and Health (NIOSH) conducts research on occupational 

diseases and injuries, responds to requests for assistance by investigating problems of health and 

safety in the workplace, recommends standards to the Occupational Safety and Health 

Administration (OSHA) and the Mine Safety and Health Administration (MSHA), and trains 

professionals in occupational safety and health. Contact: NIOSH, 200 Independence Avenue, 

SW, Washington, DC 20201 • Phone: 800-356-4674 or NIOSH Technical Information Branch, 

Robert A. Taft Laboratory, Mailstop C-19, 4676 Columbia Parkway, Cincinnati, OH 45226-1998 

• Phone: 800-35-NIOSH. 

The National Institute of Environmental Health Sciences (NIEHS) is the principal federal agency for 

biomedical research on the effects of chemical, physical, and biologic environmental agents on 

human health and well-being. Contact: NIEHS, PO Box 12233, 104 T.W. Alexander Drive, 

Research Triangle Park, NC 27709 • Phone: 919-541-3212. 

Referrals 

The Association of Occupational and Environmental Clinics (AOEC) has developed a network of clinics 

in the United States to provide expertise in occupational and environmental issues. Contact: 

AOEC, 1010 Vermont Avenue, NW, #513, Washington, DC 20005 • Phone: 202-347-4976 

• FAX: 202-347-4950 • e-mail: AOEC@AOEC.ORG • Web Page: http://www.aoec.org/. 

The American College of Occupational and Environmental Medicine (ACOEM) is an association of 

physicians and other health care providers specializing in the field of occupational and 

environmental medicine. Contact: ACOEM, 55 West Seegers Road, Arlington Heights, 

IL 60005 • Phone: 847-818-1800 • FAX: 847-818-9266. 
PBBs and PBDEs ix 

CONTRIBUTORS 

CHEMICAL MANAGER(S)/AUTHOR(S): 

Hana R. Pohl, M.D., Ph.D. 

ATSDR, Division of Toxicology, Atlanta, GA 

Stephen Bosch, B.S. 

Syracuse Research Corporation, Syracuse, NY 

Richard J. Amata, M.S. 

Syracuse Research Corporation, Syracuse, NY 

Carol J. Eisenmann, Ph.D. 

Sciences International, Inc., Alexandria, VA 

THE PROFILE HAS UNDERGONE THE FOLLOWING ATSDR INTERNAL REVIEWS: 

1. Health Effects Review. The Health Effects Review Committee examines the health effects 

chapter of each profile for consistency and accuracy in interpreting health effects and classifying 

end points. 

2. Minimal Risk Level Review. The Minimal Risk Level Workgroup considers issues relevant to 

substance-specific Minimal Risk Levels (MRLs), reviews the health effects database of each 

profile, and makes recommendations for derivation of MRLs. 

3. Data Needs Review. The Research Implementation Branch reviews data needs sections to assure 

consistency across profiles and adherence to instructions in the Guidance. 

PBBs and PBDEs xi 

PEER REVIEW 

A peer review panel was assembled for polybrominated biphenyls and polybrominated diphenyl ethers. 

The panel consisted of the following members: 

 1. Martin Alexander, Ph.D., Professor, Department of Soil and Crop Sciences, Cornell University, 

Ithaca, New York; 

 2. Loren Koller, DVM, Ph.D., College of Veterinary Medicine, Oregon State University, 105 

Magruder Hall, Corvallis, Oregon; 

 3. Christopher Metcalfe, Ph.D., Professor, Environmental and Resource Studies, Trent University, 

1600 West Bank Drive, Peterborough, Ontario, Canada; 

 4. Larry W. Robertson, Ph.D., Professor, Graduate Center for Toxicology, University of Kentucky, 

Lexington, Kentucky; and 

 5. Lee R. Shull, Ph.D., Corporate Toxicology and Risk Assessment Practice Director, Montgomery, 

Watson, and Harza, Sacramento, California. 

These experts collectively have knowledge of polybrominated biphenyls and polybrominated diphenyl 

ether's physical and chemical properties, toxicokinetics, key health end points, mechanisms of action, 

human and animal exposure, and quantification of risk to humans. All reviewers were selected in 

conformity with the conditions for peer review specified in Section 104(I)(13) of the Comprehensive 

Environmental Response, Compensation, and Liability Act, as amended. 

Scientists from the Agency for Toxic Substances and Disease Registry (ATSDR) have reviewed the peer 

reviewers' comments and determined which comments will be included in the profile. A listing of the 

peer reviewers' comments not incorporated in the profile, with a brief explanation of the rationale for their 

exclusion, exists as part of the administrative record for this compound. A list of databases reviewed and 

a list of unpublished documents cited are also included in the administrative record. 

The citation of the peer review panel should not be understood to imply its approval of the profile's final 

content. The responsibility for the content of this profile lies with the ATSDR. 

PBBs and PBDEs xiii 

CONTENTS 

DISCLAIMER ..............................................................................................................................................ii 

UPDATE STATEMENT .............................................................................................................................iii 

FOREWORD ................................................................................................................................................v 

QUICK REFERENCE FOR HEALTH CARE PROVIDERS....................................................................vii 

CONTRIBUTORS.......................................................................................................................................ix 

PEER REVIEW ...........................................................................................................................................xi 

CONTENTS...............................................................................................................................................xiii 

LIST OF FIGURES ..................................................................................................................................xvii 

LIST OF TABLES.....................................................................................................................................xix 

1. PUBLIC HEALTH STATEMENT—PBBs.............................................................................................1 

1.1 WHAT ARE PBBs? ................................................................................................................................1 

1.2 WHAT HAPPENS TO PBBs WHEN THEY ENTER THE ENVIRONMENT?..........................2 

1.3 HOW MIGHT I BE EXPOSED TO PBBs? ........................................................................................3 

1.4 HOW CAN PBBs ENTER AND LEAVE MY BODY?....................................................................3 

1.5 HOW CAN PBBs AFFECT MY HEALTH? ......................................................................................4 

1.6 HOW CAN PBBs AFFECT CHILDREN?..........................................................................................6 

1.7 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO PBBs?.................................7 

1.8 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN EXPOSED 

TO PBBs? .................................................................................................................................................7 

1.9 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO 

PROTECT HUMAN HEALTH? ..........................................................................................................8 

1.10 WHERE CAN I GET MORE INFORMATION?...............................................................................9 

2. PUBLIC HEALTH STATEMENT—PBDEs........................................................................................11 

2.1 WHAT ARE PBDEs? ...........................................................................................................................11 

2.2 WHAT HAPPENS TO PBDEs WHEN THEY ENTER THE ENVIRONMENT?.....................12 

2.3 HOW MIGHT I BE EXPOSED TO PBDEs? ...................................................................................13 

2.4 HOW CAN PBDEs ENTER AND LEAVE MY BODY?...............................................................14 

2.5 HOW CAN PBDEs AFFECT MY HEALTH? .................................................................................15 

2.6 HOW CAN PBDEs AFFECT CHILDREN?.....................................................................................16 

2.7 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO PBDEs?............................17 

2.8 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN EXPOSED 

TO PBDEs? ............................................................................................................................................18 

2.9 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO 

PROTECT HUMAN HEALTH? ........................................................................................................19 

2.10 WHERE CAN I GET MORE INFORMATION?.............................................................................20 

3. RELEVANCE TO PUBLIC HEALTH—PBBs ....................................................................................23 

3.1 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO PBBs IN THE UNITED 

STATES..................................................................................................................................................23 

3.2 SUMMARY OF HEALTH EFFECTS...............................................................................................25 

3.3 MINIMAL RISK LEVELS..................................................................................................................33 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs..................................................................................37 

4.1 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO PBDEs IN THE UNITED 

STATES..................................................................................................................................................37 
PBBs and PBDEs xiv 

4.2 SUMMARY OF HEALTH EFFECTS...............................................................................................39 

4.3 MINIMAL RISK LEVELS (MRLs)...................................................................................................45 

5. HEALTH EFFECTS..............................................................................................................................57 

5.1 INTRODUCTION ........................................................................................................................................57 

5.2 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE..............................................60 

5.2.1 Inhalation Exposure...................................................................................................................62 

5.2.1.1 Death ...................................................................................................................................62 

5.2.1.2 Systemic Effects..................................................................................................................69 

5.2.1.3 Immunological and Lymphoreticular Effects .....................................................................77 

5.2.1.4 Neurological Effects............................................................................................................78 

5.2.1.5 Reproductive Effects...........................................................................................................79 

5.2.1.6 Developmental Effects........................................................................................................80 

5.2.1.7 Cancer .................................................................................................................................80 

5.2.2 Oral Exposure ............................................................................................................................80 

5.2.2.1 Death ...................................................................................................................................80 

5.2.2.2 Systemic Effects................................................................................................................166 

5.2.2.3 Immunological and Lymphoreticular Effects ...................................................................194 

5.2.2.4 Neurological Effects..........................................................................................................199 

5.2.2.5 Reproductive Effects.........................................................................................................205 

5.2.2.6 Developmental Effects......................................................................................................208 

5.2.2.7 Cancer ...............................................................................................................................216 

5.2.3 Dermal Exposure .....................................................................................................................226 

5.2.3.1 Death .................................................................................................................................226 

5.2.3.2 Systemic Effects................................................................................................................229 

5.2.3.3 Immunological and Lymphoreticular Effects ...................................................................235 

5.2.3.4 Neurological Effects..........................................................................................................235 

5.2.3.5 Reproductive Effects.........................................................................................................236 

5.2.3.6 Developmental Effects......................................................................................................236 

5.2.3.7 Cancer ...............................................................................................................................236 

5.3 GENOTOXICITY.......................................................................................................................................237 

5.4 TOXICOKINETICS...................................................................................................................................241 

5.4.1 Absorption ...............................................................................................................................242 

5.4.1.1 Inhalation Exposure ..........................................................................................................242 

5.4.1.2 Oral Exposure ...................................................................................................................243 

5.4.1.3 Dermal Exposure...............................................................................................................245 

5.4.2 Distribution..............................................................................................................................246 

5.4.2.1 Inhalation Exposure ..........................................................................................................246 

5.4.2.2 Oral Exposure ...................................................................................................................247 

5.4.2.3 Dermal Exposure...............................................................................................................253 

5.4.2.4 Other Routes of Exposure .................................................................................................254 

5.4.3 Metabolism ..............................................................................................................................255 

5.4.4 Elimination and Excretion .......................................................................................................261 

5.4.4.1 Inhalation Exposure ..........................................................................................................261 

5.4.4.2 Oral Exposure ...................................................................................................................262 

5.4.4.3 Dermal Exposure...............................................................................................................266 

5.4.4.4 Other Routes of Exposure .................................................................................................266 

5.4.5 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) Models................266 

5.5 MECHANISMS OF ACTION..................................................................................................................269 

5.5.1 Pharmacokinetic Mechanisms .................................................................................................269 

5.5.2 Mechanisms of Toxicity ..........................................................................................................270 
PBBs and PBDEs xv 

5.5.3 Animal-to-Human Extrapolations............................................................................................284 

5.6 TOXICITIES MEDIATED THROUGH THE NEUROENDOCRINE AXIS...................................285 

5.7 CHILDREN’S SUSCEPTIBILITY..........................................................................................................291 

5.8 BIOMARKERS OF EXPOSURE AND EFFECT.................................................................................303 

5.8.1 Biomarkers Used to Identify or Quantify Exposure to PBBs and PBDEs...............................304 

5.8.2 Biomarkers Used to Characterize Effects Caused by PBBs and PBDEs.................................307 

5.9 INTERACTIONS WITH OTHER CHEMICALS .................................................................................310 

5.10 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE........................................................311 

5.11 METHODS FOR REDUCING TOXIC EFFECTS .............................................................................313 

5.11.1 Reducing Peak Absorption Following Exposure...................................................................314 

5.11.2 Reducing Body Burden..........................................................................................................314 

5.11.3 Interfering with the Mechanism of Action for Toxic Effects ................................................315 

5.12 ADEQUACY OF THE DATABASE....................................................................................................316 

5.12.1 Existing Information on Health Effects of PBBs and PBDEs ...............................................317 

5.12.2 Identification of Data Needs..................................................................................................320 

5.12.3 Ongoing Studies.....................................................................................................................340 

6. CHEMICAL AND PHYSICAL INFORMATION..............................................................................345 

6.1 CHEMICAL IDENTITY....................................................................................................................345 

6.2 PHYSICAL AND CHEMICAL PROPERTIES .............................................................................355 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL..........................................................367 

7.1 PRODUCTION....................................................................................................................................367 

7.2 IMPORT/EXPORT .............................................................................................................................370 

7.3 USE........................................................................................................................................................370 

7.4 DISPOSAL...........................................................................................................................................374 

8. POTENTIAL FOR HUMAN EXPOSURE .........................................................................................379 

8.1 OVERVIEW.........................................................................................................................................379 

8.2 RELEASES TO THE ENVIRONMENT.........................................................................................383 

8.2.1 Air .......................................................................................................................................387 

8.2.2 Water...................................................................................................................................392 

8.2.3 Soil ......................................................................................................................................394 

8.3 ENVIRONMENTAL FATE..............................................................................................................395 

8.3.1 Transport and Partitioning...................................................................................................395 

8.3.2 Transformation and Degradation ........................................................................................401 

8.3.2.1 Air...............................................................................................................................401 

8.3.2.2 Water ..........................................................................................................................402 

8.3.2.3 Sediment and Soil.......................................................................................................407 

8.3.2.4 Other Media................................................................................................................409 

8.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT...................................409 

8.4.1 Air .......................................................................................................................................409 

8.4.2 Water...................................................................................................................................412 

8.4.3 Sediment and Soil ...............................................................................................................412 

8.4.4 Other Environmental Media................................................................................................416 

8.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE..........................................450 

8.6 EXPOSURES OF CHILDREN.........................................................................................................454 

8.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES.................................................456 

8.8 ADEQUACY OF THE DATABASE...............................................................................................457 

8.8.1 Identification of Data Needs ...............................................................................................457 

8.8.2 Ongoing Studies..................................................................................................................465 
PBBs and PBDEs xvi 

9. ANALYTICAL METHODS ...............................................................................................................467 

9.1 BIOLOGICAL MATERIALS ...........................................................................................................469 

9.2 ENVIRONMENTAL SAMPLES .....................................................................................................473 

9.3 ADEQUACY OF THE DATABASE...............................................................................................476 

9.3.1 Identification of Data Needs ...............................................................................................477 

9.3.2 Ongoing Studies..................................................................................................................478 

10. REGULATIONS AND ADVISORIES .............................................................................................479

11. REFERENCES ..................................................................................................................................487 

12. GLOSSARY ......................................................................................................................................559 

APPENDICES 

A. ATSDR MINIMAL RISK LEVELS AND WORKSHEETS .............................................................A-1 

B. USER’S GUIDE.................................................................................................................................. B-1 

C. ACRONYMS, ABBREVIATIONS, AND SYMBOLS...................................................................... C-1 

D. INDEX ................................................................................................................................................D-1 
PBBs and PBDEs xvii 

LIST OF FIGURES 

5-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ether—Inhalation..................... 67 

5-2. Levels of Significant Exposure to Polybrominated Biphenyls—Oral ......................................... 129 

5-3. Levels of Significant Exposure to Lower Brominated Diphenyl Ether—Oral ............................ 147 

5.4. Levels of Significant Exposure to Decabromodiphenyl Ether—Oral.......................................... 160 

5-5. Conceptual Representation of a Physiologically Based Pharmacokinetic (PBPK) Model for a 

Hypothetical Chemical Substance ............................................................................................... 268 

5-6. Existing Information on Health Effects of PBBs......................................................................... 318

5-7. Existing Information on Health Effects of PBDEs ...................................................................... 319 

8-1. Frequency of NPL Sites with PBB Contamination...................................................................... 382 

PBBs and PBDEs xix 

LIST OF TABLES 

5-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ether—Inhalation..................... 63 

5-2. Levels of Significant Exposure to Polybrominated Biphenyls—Oral ........................................... 81 

5-3. Levels of Significant Exposure to Lower Brominated Diphenyl Ether—Oral ............................ 136 

5-4. Levels of Significant Exposure to Decabromodiphenyl Ether—Oral.......................................... 150 

5-5. Levels of Significant Exposure to Polybrominated Biphenyls—Dermal .................................... 227 

5-6. Levels of Significant Exposure to Polybrominated Biphenyl Ethers—Dermal........................... 230 

5-7. Genotoxicity of PBBs and PBDEs In Vivo .................................................................................. 238 

5-8. Genotoxicity of PBBs and PBDEs In Vitro ................................................................................. 239 

5-9. Ongoing Studies on the Health Effects of PBBs and PBDEs ..................................................... 341 

6-1. Chemical Identity of Polybrominated Biphenyl (PBB) and Polybrominated Diphenyl Ether 

(PBDE) Congeners....................................................................................................................... 347 

6-2. Chemical Identity of Selected PBBs............................................................................................ 353 

6-3. Chemical Identity of Technical PBDEs....................................................................................... 356 

6-4. Physical and Chemical Properties of Selected PBBs................................................................... 357 

6-5. Identified PBB Congeners in FireMaster® BP-6 and FireMaster® FF-1.................................... 359 

6-6. Physical and Chemical Properties of Some Technical PBDE Mixtures ...................................... 362 

6-7. Physical and Chemical Properties of Some PBDE Congeners .................................................... 364 

6-8. Composition of Commercial Brominated Diphenyl Ethers......................................................... 366 

7-1. Total Market Demand of PBDEs by Region of the World in 2001............................................. 371 

7-2. Facilities that Produce, Process, or Use Decabromodiphenyl Ether............................................ 373 

8-1. Releases to the Environment from Facilities that Produce, Process, or Use Decabromobiphenyl 

Ether............................................................................................................................................. 392 

8-2. Concentrations (pg/m3

) of Several PBDEs in Air Samples ......................................................... 413 

8-3. Concentrations (ng/g) of Several PBDEs in Sediment and Suspended Particulate Samples....... 416 
PBBs and PBDEs xx 

8-4. Mean Concentration of Nine PBB Congeners in Lake Trout from the Great Lakes (pg/g Wet 

Weight) ........................................................................................................................................ 420 

8-5. Tissue Levels of PBBs in Michigan Residents ............................................................................ 424

8-6. Concentrations (ng/g Dry Weight) of Several PBDEs in Biosolids (Sewage Sludge) and 

 Effluent Samples.......................................................................................................................... 428 

8-7. Concentrations (ng/g Lipid Weight, Except as Noted) of Several PBDEs in Freshwater Fish 

Samples from the United States................................................................................................... 430 

8-8. Concentrations (ng/g Lipid Weight, Except as Noted) of Several PBDEs in Freshwater Fish 

Samples from the Europe............................................................................................................. 433 

8-9. Concentrations (ng/g Lipid Weight) of Several PBDEs in Marine Aquatic Species................... 436 

8-10. Concentrations (ng/g Lipid Weight) of Several PBDEs in Marine Animals ............................... 440 

8-11. Concentrations (ng/g Lipid Weight) of Several PBDEs in Human Blood Samples .................... 445 

8-12. Concentrations (ng/g Lipid Weight) of Several PBDEs in Human Adipose Tissue Samples ..... 446 

8-13. Concentrations (ng/g Lipid Weight) of Several PBDEs in Human Breast Milk Samples........... 447 

9-1. Analytical Methods for Determining PBBs in Biological Materials ........................................... 473 

9-2. Analytical Methods for Determining PBDEs in Biological Materials......................................... 474 

9-3. Analytical Methods for Determining PBBs in Environmental Samples...................................... 476 

9-4. Analytical Methods for Determining PBDEs in Environmental Samples ................................... 477 

10-1. Regulations and Guidelines Applicable to PBBs and PBDEs ..................................................... 482 
PBBs AND PBDEs 1 

1. PUBLIC HEALTH STATEMENT—PBBs 

This public health statement tells you about polybrominated biphenyls (PBBs) and the effects of 

exposure to PBBs. 

The Environmental Protection Agency (EPA) identifies the most serious hazardous waste sites in 

the nation. These sites are then placed on the National Priorities List (NPL) and are targeted for 

long-term federal clean-up activities. PBBs have been found in at least nine of the 1,647 current 

or former NPL sites. Although the total number of NPL sites evaluated for these substances is 

not known, the possibility exists that the number of sites at which PBBs are found may increase 

in the future as more sites are evaluated. This information is important because these sites may 

be sources of exposure and exposure to these substances may harm you. 

When a substance is released either from a large area, such as an industrial plant, or from a 

container, such as a drum or bottle, it enters the environment. Such a release does not always 

lead to exposure. You can be exposed to a substance only when you come in contact with it. 

You may be exposed by breathing, eating, or drinking the substance, or by skin contact. 

If you are exposed to PBBs, many factors will determine whether you will be harmed. These 

factors include the dose (how much), the duration (how long), and how you come in contact with 

PBBs. You must also consider any other chemicals you are exposed to and your age, sex, diet, 

family traits, lifestyle, and state of health. 

1.1 WHAT ARE PBBs? 

Polybrominated biphenyls (PBBs) are chemicals that were added to plastics used in a variety of 

consumer products, such as computer monitors, televisions, textiles, and plastic foams, to make 

them difficult to burn. Because PBBs were mixed into plastics rather than bound to them, they 

were able to leave the plastic and find their way into the environment. Commercial production 

of PBBs began in the 1970s. Manufacture of PBBs was discontinued in the United States in 
PBBs AND PBDEs 2 

1. PUBLIC HEALTH STATEMENT—PBBs 

1976. Concern regarding PBBs is mainly related to exposures resulting from an agriculture 

contamination episode that occurred in Michigan over a 10-month period during 1973–1974. 

There are no known natural sources of PBBs in the environment. PBBs are solids and are 

colorless to off-white. PBBs enter the environment as mixtures containing a variety of 

individual brominated biphenyl (for PBBs) components, known as congeners. Some commercial 

PBB mixtures are known in the United States under the industrial trade name, FireMaster®. 

However, other flame retardant chemicals also may be identified by this name. PBBs are no 

longer used in North America because the agriculture contamination episode that occurred in 

Michigan in 1973–1974 led to the cessation of its production. More information on the physical 

properties and uses of PBBs can be found in Chapters 6 and 7. 

1.2 WHAT HAPPENS TO PBBs WHEN THEY ENTER THE ENVIRONMENT? 

In the past, PBBs entered the air, water, and soil during their manufacture and use. In addition, 

animal feed was accidentally mixed with 500–1,000 pounds of PBBs in lower Michigan in 1973. 

This contamination of the food chain affected millions of farm animals and humans living in 

Michigan at this time. PBBs entered the environment during the disposal of contaminated 

animal feed and animal products during the agriculture contamination episode. PBBs also 

entered the environment from PBB-containing wastes that manufacturers disposed of in waste 

sites. Small quantities of PBBs also entered the environment from accidental spills during 

transport. PBBs are no longer manufactured in North America, but very small amounts of PBBs 

may be released into the environment from poorly maintained hazardous waste sites and 

improper incineration of plastics that contain PBBs. 

More information about what happens to PBBs in the environment can be found in Chapter 8. 
PBBs AND PBDEs 3 

1. PUBLIC HEALTH STATEMENT—PBBs 

1.3 HOW MIGHT I BE EXPOSED TO PBBs? 

PBBs are no longer produced or used in the United States. Thus, the general population 

exposure to PBBs will only be from past releases. For people living in the lower peninsula of 

Michigan, especially near PBB contaminated areas, exposure to PBBs may still be occurring 

today. However, environmental levels have decreased since the 1970s and current exposure, if 

any, will be at low levels. For other regions of the United States, the levels of exposure will 

either be very low or none. 

Measured data in air, water, soil, and food, as well as body burden data (blood, urine, breast 

milk, and body fat), indicate that most people within the state of Michigan who were exposed to 

PBBs received very low levels of PBBs. The levels from exposure were slightly higher for 

people living in the lower peninsula of Michigan and highest among people living on 

contaminated dairy farms. Consumption of contaminated meat and dairy products caused the 

higher levels of PBBs in the body. Monitoring of the workplace environment, as well as the 

blood, urine, and body fat of workers indicated that those in PBB industries were exposed to 

higher levels of PBBs than the general population. These workers were exposed to PBBs by 

breathing contaminated workplace air and by skin contact with PBBs. Occupational exposure 

also could have occurred from the incineration of materials containing PBBs. Exposure in 

workplaces is no longer likely because PBBs are no longer manufactured. People who live near 

hazardous waste sites that contain PBBs may be exposed primarily by breathing air that contains 

PBBs. 

More information about exposure to PBBs can be found in Chapter 8. 

1.4 HOW CAN PBBs ENTER AND LEAVE MY BODY? 

If you breathe air that contains PBBs, or swallow food, water, or soil contaminated with PBBs, 

they can enter your body through your lungs and stomach and pass into the bloodstream. We 

don’t know how much of the PBBs will pass into the blood from the lungs; although most will 

probably pass into the blood from the stomach and intestines. If you touch soil containing PBBs, 
PBBs AND PBDEs 4 

1. PUBLIC HEALTH STATEMENT—PBBs 

it is highly unlikely that PBBs would pass through your skin into the bloodstream. It is not 

known how fast PBBs enter the blood from the lungs or stomach. There are no known current 

sources of PBBs because they are no longer produced and used in North America. They are 

rarely found in air and drinking water away from production plants and contaminated sites. 

Once PBBs are in your body, they can partially change into breakdown products called 

metabolites. Some metabolites and unchanged PBBs could leave your body, mainly in the feces 

and in very small amounts in the urine, within a few days. Other unchanged PBBs might stay in 

your body for many years. PBBs are stored mainly in your body fat, tend to concentrate in breast 

milk fat, and can enter the bodies of children through breast feeding. PBBs also can enter the 

bodies of unborn babies through the placenta. More information on how PBBs can enter and 

leave your body can be found in Chapter 5. 

1.5 HOW CAN PBBs AFFECT MY HEALTH? 

Scientists use many tests to protect the public from harmful effects of toxic chemicals and to find 

ways for treating persons who have been harmed. 

One way to learn whether a chemical will harm people is to determine how the body absorbs, 

uses, and releases the chemical. That sometimes involves animal testing. Animal testing may 

help identify health effects such as cancer or birth defects. Without laboratory animals, scientists 

would lose a basic method for getting information needed to make wise decisions that protect 

public health. Scientists have the responsibility to treat research animals with care and 

compassion. Scientists must comply with strict animal care guidelines because laws today 

protect the welfare of research animals. 

Much of what is known about the health effects of PBBs in people comes from studies of 

ingestion in Michigan in the early-to-mid 1970s, where feed for farm animals was accidentally 

contaminated with a fire retardant containing PBBs. People were exposed to PBBs for several 

months when they ate meat, milk, and eggs from the contaminated animals. After news of the 

contamination episode became widespread, many Michigan residents complained of various 

health problems, including nausea, abdominal pain, loss of appetite, joint pain, fatigue, and 
PBBs AND PBDEs 5 

1. PUBLIC HEALTH STATEMENT—PBBs 

weakness. However, it could not be clearly established that any of the problems were caused by 

eating the food contaminated with PBBs. PBBs also did not cause any definite changes in the 

livers or immune systems of the Michigan residents. However, some people who ate the 

contaminated food developed skin disorders, such as acne and hair loss. It is likely that PBBs 

caused the skin problems because other chemicals similar to PBBs also cause these effects. 

Workers who were exposed to PBBs for a few days to months by breathing and skin contact also 

developed acne, although not all persons exposed to PBBs developed acne. Very little is known 

about the health of people who are exposed to low levels of PBBs for long periods by eating, 

breathing, or skin contact. 

Laboratory animals fed PBBs had body weight loss, skin disorders, and nervous system effects, 

and their livers, kidneys, thyroid glands, and immune systems were seriously injured. Some 

animals fed high amounts died. PBBs also caused birth defects in animals, but it is not known 

for sure whether PBBs make males or females infertile. Most of the effects in animals occurred 

after they ate large amounts of PBBs for short periods or smaller amounts for several weeks or 

months. Body weight loss and effects on the livers, kidneys, and thyroid glands were observed. 

A lifetime study of rats and mice fed PBBs at doses higher than those expected from 

environmental exposure. A few studies tested animals exposed to PBBs by skin contact. These 

animals had injuries to the liver and skin. Only one study tested animals exposed to PBBs by 

breathing, and no health effects were observed. 

We do not know if PBBs caused or will cause cancer in people who ate food contaminated with 

PBBs. Rats developed cancer in their livers after eating a large amount of a PBB mixture only 

once. The babies of exposed rats developed cancer in their livers after eating a large amount of 

the same PBB mixture only once. Liver cancer also developed in rats and mice that ate smaller 

amounts of the PBB mixture for several months. Mice that had skin contact with a small amount 

of a PBB mixture for several months did not develop skin cancer. There are no cancer studies in 

animals that breathed PBBs. Because animals fed PBBs did develop cancer, the National 

Toxicology Program (NTP) of the Department of Health and Human Services (DHHS) 

determined that PBBs may reasonably be anticipated to be carcinogens. Similarly, the 
PBBs AND PBDEs 6 

1. PUBLIC HEALTH STATEMENT—PBBs 

International Agency for Research on Cancer (IARC) has determined that PBBs are possibly 

carcinogenic to humans. The EPA has not classified the carcinogenicity of PBBs. 

We do not know whether the effects found in animals exposed to PBBs would also occur in 

people exposed in the same way. The amounts of PBBs that caused health effects in animals are 

much greater than levels of PBBs normally found in the environment. Long-term exposure to 

these chemicals has a greater potential to cause health effects than does short-term exposure to 

low levels because they tend to build up in your body over many years. More information on 

how PBBs can affect your health can be found in Chapter 5. 

1.6 HOW CAN PBBs AFFECT CHILDREN? 

This section discusses potential health effects in humans from exposures during the period from 

conception to maturity at 18 years of age. 

Children are exposed to PBBs in generally the same way as are adults, mainly by eating 

contaminated food. Because of their smaller weight, children’s intake of PBBs per kilogram (or 

pound) of body weight may be greater than that of adults. The most likely way that infants will 

be exposed is from breast milk that contains PBBs, although fetuses in the womb are also 

exposed. Children who live near hazardous waste sites might accidentally eat some PBBs by 

putting dirty hands or other soil/dirt-covered objects in their mouths, by eating without washing 

their hands, or similar behavior. Some children also eat dirt on purpose. It is not possible that 

children could be exposed to PBBs following transport of the chemical on clothing from the 

parent’s workplace to the home because PBBs are no longer being produced or used. 

Some information on health effects of PBBs in children is available from studies of the Michigan 

contamination episode. Symptoms of ill health were not associated with increased exposure to 

PBBs, and general neurological examinations did not show any abnormalities. More detailed 

studies of physical and neuropsychological development showed no effects that were clearly 

related to PBBs among Michigan children exposed during the episode. Changes in nerve and 

brain function have been seen in animals that were exposed to PBBs in the womb and by 
PBBs AND PBDEs 7 

1. PUBLIC HEALTH STATEMENT—PBBs 

nursing. Animal studies also found that exposure to PBBs during pregnancy or lactation caused 

changes in thyroid hormone levels in the newborn animals and, at high doses, increases in 

prenatal death and structural birth defects. 

As indicated above, children can be exposed to PBBs before birth and from breast milk. PBBs 

are stored in the mother’s body and can be released during pregnancy, cross the placenta, and 

enter fetal tissues. Because PBBs dissolve readily in fat, they can accumulate in breast milk fat 

and be transferred to babies and young children. PBBs have been found in breast milk; however, 

in most cases, the benefits of breast-feeding outweigh any risks from exposure in mother’s milk. 

You should consult your health care provider if you have any concerns about PBBs and breast 

feeding. Because the nervous system and thyroid are still developing in the fetus and child, the 

effects of PBBs on these target systems might be more profound from exposure before and soon 

after birth. That could mean fetuses and children are more susceptible to PBBs than are adults. 

1.7 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO PBBs? 

If your doctor finds that you have been exposed to substantial amounts of PBBs, ask whether 

your children might also have been exposed. Your doctor might need to ask your state health 

department to investigate. 

Because PBBs are no longer produced or used, the risk of exposure to these compounds is 

limited. You and your children could be exposed to PBBs by eating fish or wildlife caught from 

contaminated locations. Children who live near hazardous waste sites should be discouraged 

from playing in the dirt near these sites because they could still contain PBBs. Children should 

also be discouraged from eating the dirt, and careful handwashing practices should be followed. 

1.8 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN 

EXPOSED TO PBBs? 

Special tests can determine whether PBBs are in the blood, body fat, and breast milk. These are 

not regular or routine clinical tests, but could be ordered by a doctor to detect PBBs in people 
PBBs AND PBDEs 8 

1. PUBLIC HEALTH STATEMENT—PBBs 

exposed to them in the environment and at work. If your PBB levels are higher than the normal 

levels, this will show that you have been exposed to high levels of the chemicals. However, 

these measurements cannot determine the exact amount or type of PBBs that you have been 

exposed to, or how long you have been exposed. Although these tests can indicate whether you 

have been exposed to PBBs to a greater extent than the general population, they do not predict 

whether you will be harmed. Blood tests are the easiest, safest, and probably the best method for 

detecting recent or past exposures to large amounts of PBBs. Results of such tests should be 

reviewed and carefully interpreted by physicians with a background in environmental and 

occupational medicine. Exposures to PBBs have been of greatest concern in Michigan as 

explained in Sections 1.3 and 1.5. More information on tests used to determine whether you 

have been exposed to PBBs can be found in Chapter 5 (Section 5.11) and Chapter 9 

(Section 9.1). 

1.9 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO 

PROTECT HUMAN HEALTH? 

The federal government develops regulations and recommendations to protect public health. 

Regulations can be enforced by law. The EPA, the Occupational Safety and Health 

Administration (OSHA), and the Food and Drug Administration (FDA) are some federal 

agencies that develop regulations for toxic substances. Recommendations provide valuable 

guidelines to protect public health, but cannot be enforced by law. The Agency for Toxic 

Substances and Disease Registry (ATSDR) and the National Institute for Occupational Safety 

and Health (NIOSH) are two federal organizations that develop recommendations for toxic 

substances. 

Regulations and recommendations can be expressed as “not-to-exceed” levels, that is, levels of a 

toxic substance in air, water, soil, or food that do not exceed a critical value that is usually based 

on levels that affect animals; they are then adjusted to levels that will help protect humans. 

Sometimes these not-to-exceed levels differ among federal organizations because they used 

different exposure times (an 8-hour workday or a 24-hour day), different animal studies, or other 

factors. 
PBBs AND PBDEs 9 

1. PUBLIC HEALTH STATEMENT—PBBs 

Recommendations and regulations are also updated periodically as more information becomes 

available. For the most current information, check with the federal agency or organization that 

provides it. Some regulations and recommendations for PBBs include the following: 

At present, there are no federal guidelines or recommendations for protecting human health from 

exposure to PBBs. 

1.10 WHERE CAN I GET MORE INFORMATION? 

If you have any more questions or concerns about PBBs, please contact your community or state 

health or environmental quality department, or contact ATSDR at the address and phone number 

below. 

ATSDR can also tell you the location of occupational and environmental health clinics. These 

clinics specialize in recognizing, evaluating, and treating illnesses that result from exposure to 

hazardous substances. 

Toxicological profiles are also available on-line at www.atsdr.cdc.gov/toxpro2.html and on 

CD-ROM. You may request a copy of the ATSDR ToxProfilesTM CD-ROM by calling the tollfree information and technical assistance number at 1-888-42ATSDR (1-888-422-8737), by 

e-mail at atsdric@cdc.gov, or by writing to: 

 Agency for Toxic Substances and Disease Registry 

 Division of Toxicology 

 1600 Clifton Road NE 

 Mailstop F-32 

 Atlanta, GA 30333 

 Fax: 1-770-488-4178 
PBBs AND PBDEs 10 

1. PUBLIC HEALTH STATEMENT—PBBs 

Organizations for-profit may request copies of final Toxicological Profiles from the following: 

 National Technical Information Service (NTIS) 

 5285 Port Royal Road 

 Springfield, VA 22161 

 Phone: 1-800-553-6847 or 1-703-605-6000 

 Web site: http://www.ntis.gov/ 
PBBs and PBDEs 11 

2. PUBLIC HEALTH STATEMENT—PBDEs 

This public health statement tells you about polybrominated diphenyl ethers (PBDEs) and the 

effects of exposure to PBDEs. 

The Environmental Protection Agency (EPA) identifies the most serious hazardous waste sites in 

the nation. These sites are then placed on the National Priorities List (NPL) and are targeted for 

long-term federal clean-up activities. PBDEs have not been found in the 1,647 current or former 

NPL sites. Although the total number of NPL sites evaluated for these substances is not known, 

the possibility exists that PBDEs may be found in the future as more sites are evaluated. This 

information is important because these sites may be sources of exposure and exposure to these 

substances may harm you. 

When a substance is released either from a large area, such as an industrial plant, or from a 

container, such as a drum or bottle, it enters the environment. Such a release does not always 

lead to exposure. You can be exposed to a substance only when you come in contact with it. 

You may be exposed by breathing, eating, or drinking the substance, or by skin contact. 

If you are exposed to PBDEs, many factors will determine whether you will be harmed. These 

factors include the dose (how much), the duration (how long), and how you come in contact with 

PBDEs. You must also consider any other chemicals you are exposed to and your age, sex, diet, 

family traits, lifestyle, and state of health. 

2.1 WHAT ARE PBDEs? 

PBDEs are flame-retardant chemicals that are added to a variety of consumer products to make 

them difficult to burn. Because PBDEs are added rather than reacted to the product, they could 

leave the product under ideal conditions and enter the environment, but this rarely happens. The 

first commercial productions of PBDEs began in the 1970s in Germany. Production of PBDEs 

has continued until the present. There are three commercial PBDE products (i.e., penta-, octa-, 
PBBs and PBDEs 12 

2. PUBLIC HEALTH STATEMENT—PBDEs 

and decabromodiphenyl ethers). Deca- and octa-brominated types of PBDEs are also produced 

outside of the United States (in China and Israel). Decabromodiphenyl ether (decaBDE) makes 

up 82% of these products manufactured globally. Its main use is for electronic enclosures, such 

as television cabinets. Octabromodiphenyl ether (octaBDE) product is used in plastics for 

business equipment. Pentabromodiphenyl ether (pentaBDE) product is used in foam for 

cushioning in upholstery. PBDEs have not been associated with actual health-related effects. 

Concerns have increased, however, because some of these chemicals (particularly PBDEs in the 

pentaBDEs) have been found in the environment at varying concentrations. Environmental 

concentrations of lower brominated PBDEs, which may be leveling off in Europe, appear to be 

increasing in certain areas of Canada and the United States. 

PBDEs are a group of synthetic organic chemicals with no known natural sources in the 

environment, except for a few marine organisms that produce forms of PBDEs that contain 

higher levels of oxygen. Commercial decaBDE and octaBDE products are colorless to off-white 

solids, whereas commercial pentaBDE product is a thick liquid. PBDEs are not expected to 

evaporate into the air. PBDEs in the air are mostly found with dust rather than as a vapor. 

PBDEs enter the environment as mixtures containing a variety of individual brominated diphenyl 

ether (for PBDEs) components, known as congeners. Congeners are distinct members of a class 

of chemical substances. Some commercial mixtures of PBDEs may be known by their industrial 

trade names, (i.e., DE-60F Special, DE-61, DE-62, DE-71, DE-79, DE-83R, Saytex® 102E). 

PBDEs are still produced and widely used in the United States, although the sole manufacturer of 

penta- and octaPBDE commercial products in the United States is expected to quit making these 

chemicals by the end of 2004. More information on the physical properties and uses of PBDEs 

can be found in Chapters 6 and 7. 

2.2 WHAT HAPPENS TO PBDEs WHEN THEY ENTER THE ENVIRONMENT? 

PBDEs enter air, water, and soil during their manufacture and use in consumer products. When 

PBDEs are suspended in air, they can be present as particles. They eventually return to land or 

water as the dust settles and are washed out by snow and rainwater. It is not yet possible to say 

how long PBDEs remain in the air. PBDEs do not dissolve easily in water, and therefore, high 
PBBs and PBDEs 13 

2. PUBLIC HEALTH STATEMENT—PBDEs 

levels of PBDEs are not found in water. The very small amounts of PBDEs that do occur in 

water stick to particles and eventually settle to the bottom. Sediments at the bottom of bodies of 

water, such as lakes and rivers, generally act as reservoirs for decaBDEs, which can remain there 

for years. Some lower brominated PBDEs (e.g., tetra- and penta-congeners of PBDE) in water 

may build up in fish to low concentrations (about 10 billionths of a gram to 1 millionth of a gram 

of PDBE per gram of fresh fish [or 10x10-9–1x10-6 grams of PBDE per gram of fresh fish]). 

However, higher brominated PBDEs, such as decaBDE, are not found in fish at measurable 

concentrations. In general, the breakdown of PBDEs in soil is very slow, so they may remain in 

soil for several years. PBDEs bind strongly to soil particles. Rainwater is not expected to spread 

them much below the soil surface; thus, it is unlikely that PBDEs will enter groundwater. 

More information about what happens to PBDEs in the environment can be found in Chapter 8. 

2.3 HOW MIGHT I BE EXPOSED TO PBDEs? 

Certain PBDE mixtures, especially decaBDE, are produced in many places in the world. 

Currently, the United States is the only producer of the pentaBDE mixture for commerce. 

However, production pentaBDE and octaBDE for commerce soon will begin to be phased out in 

the United States and elsewhere. Lower brominated PBDEs, such as tetraBDE and pentaBDE 

congeners, are found throughout the environment and are found at low levels in air, sediments, 

animals, and food. The concentrations of lower brominated PBDEs in blood, breast milk, and 

body fat indicate that most people are exposed to low levels of these PBDEs. Concentrations of 

lower brominated PBDEs have been increasing in tissues and body fluids of individuals living in 

the United States. Currently, levels of lower brominated PBDE congeners in individuals living 

in the United States are greater than levels reported in other regions of the world. Highly 

brominated congeners, such as decaBDE, are not commonly found throughout the environment. 

They may be found at low levels close to places where they are produced or used. In 2001, 

PBDEs were detected in dust and smoke samples taken near the World Trade Center disaster 

site. Although no definitive studies in the United States have been conducted to identify the 

sources of exposure, people appear to be exposed to the lower brominated congeners of PBDEs 

by eating food that contains these PBDEs. In the United States, the concentration of PBDEs 
PBBs and PBDEs 14 

2. PUBLIC HEALTH STATEMENT—PBDEs 

(primarily the tetra- and penta-brominated congeners of PBDEs) in outdoor air ranges from 2 to 

77 trillionths of a gram per cubic meter (or 2–77x10-12 grams/m3

), which indicates low levels of 

exposure of the general population to these PBDEs. Indoor air concentrations of PBDEs in 

lecture halls, indoor environments with computers, and rooms with computers or other electronic 

devices, such as television sets, also have low levels of PBDEs in suspended dust. Workers 

involved in the manufacture and production of PBDE-containing resins are exposed to higher 

concentrations of PBDEs. Occupational exposure can also occur in confined workplaces where 

plastic and foam products containing PBDEs are recycled, or where computer monitors 

containing PBDEs are repaired. People who live near hazardous waste sites may be exposed to 

PBDEs by breathing air containing PBDE-contaminated dust. However, PBDE eventually 

settles or washes out of the air, so this potential exposure route likely to be minimal. 

More information about exposure to PBDEs can be found in Chapter 8. 

2.4 HOW CAN PBDEs ENTER AND LEAVE MY BODY? 

The main source of exposure to PBDEs may be through foods, particularly those with high fat 

content, such as fatty fish. Some lower brominated PBDEs have been detected in air samples, 

indicating that people can also be exposed by inhalation. The ways that PBDEs might enter and 

leave your body depend on the chemical structures of the congener components. The higher 

brominated PBDEs, particularly decaBDE (the major PBDE in use today), act much differently 

in the body than do lower brominated PBDEs. If you breathe air that contains PBDEs, or 

swallow food, water, or soil contaminated with PBDEs, the lower brominated congeners are 

much more likely than decaBDE to enter your body through your lungs and stomach and pass 

into the bloodstream. If you touch soil containing PBDEs, which could happen at a hazardous 

waste site, it is highly unlikely that either lower or higher brominated PBDEs would pass through 

your skin into the bloodstream. Once PBDEs are in your body, the congeners might partially 

change into breakdown products called metabolites. DecaBDE might leave your body 

unchanged or as metabolites, mainly in the feces and in very small amounts in the urine, within a 

few days. Lower brominated PBDEs, generally tetra-, penta-, and hexaBDE congeners, might 

stay in your body for many years, stored mainly in body fat. The lower brominated PBDEs also 
PBBs and PBDEs 15 

2. PUBLIC HEALTH STATEMENT—PBDEs 

tend to concentrate in breast milk fat, and can enter the bodies of children through breast feeding. 

PBDEs also can enter the bodies of unborn babies through the placenta. More information on 

how PBDEs can enter and leave your body can be found in Chapter 5. 

2.5 HOW CAN PBDEs AFFECT MY HEALTH? 

Scientists use many tests to protect the public from harmful effects of toxic chemicals and to find 

ways for treating persons who have been harmed. 

One way to learn whether a chemical will harm people is to determine how the body absorbs, 

uses, and releases the chemical. For some chemicals, animal testing may be necessary. Animal 

testing may also help identify health effects such as cancer or birth defects. Without laboratory 

animals, scientists would lose a basic method for getting information needed to make wise 

decisions that protect public health. Scientists have the responsibility to treat research animals 

with care and compassion. Scientists must comply with strict animal care guidelines because 

laws today protect the welfare of research animals. 

Nothing definite is known about the health effects of PBDEs in people. Practically all of the 

available information is from studies of laboratory animals. Animal studies indicate that 

commercial decaBDE mixtures are generally much less toxic than the products containing lower 

brominated PBDEs. Because of its very different toxicity, decaBDE is expected to have 

relatively little effect on the health of humans. Rats and mice that ate food containing moderate 

amounts of lower brominated PBDEs for short periods of time had mainly thyroid effects. Rats 

and mice that ate smaller amounts over several weeks or months developed effects in the liver 

and in the thyroid. It is speculated that many of the thyroid effects of PBDEs are specific to the 

species of test animals, suggesting that they are less likely to occur in humans. Subtle behavioral 

changes have been observed in animals exposed to PBDEs as infants. One possible explanation 

for the behavioral effects might be related to changes in the thyroid, because development of the 

nervous system is dependent on thyroid hormones. PBDEs have not caused other kinds of birth 

defects in animals, but more studies are needed to determine if PBDEs can impair reproduction. 

Preliminary findings from short-term animal studies suggest that some PBDEs might impair the 
PBBs and PBDEs 16 

2. PUBLIC HEALTH STATEMENT—PBDEs 

immune system. Animals exposed to PBDEs by skin contact showed signs of skin 

irritation only if they had been scratched. 

We don’t know if PBDEs can cause cancer in people, although liver tumors developed in rats 

and mice that ate extremely large amounts of decaBDE throughout their lifetime. On the basis of 

evidence for cancer in animals, decaBDE is classified as a possible human carcinogen by EPA. 

Lower brominated PBDEs have not yet been tested for cancer. Neither the U.S. Department of 

Health and Human Services (DHHS) nor the International Agency for Research on Cancer 

(IARC) have classified the carcinogenicity of any PBDEs. 

We don’t know whether the effects found in animals exposed to PBDEs would also occur in 

people exposed in the same way. The amounts of PBDEs that caused health effects in animals 

are much greater than levels of PBDEs normally found in the environment. Long-term exposure 

to PBDEs has a greater potential to cause health effects than does short-term exposure to low 

levels because of their tendency to build up in your body over many years. Additionally, the 

lower brominated commercial pentaBDE and octaBDE products are much more likely to cause 

health effects than is decaBDE. More information on how PBDEs can affect your health can be 

found in Chapter 5. 

2.6 HOW CAN PBDEs AFFECT CHILDREN? 

This section discusses potential health effects in humans from exposures during the period from 

conception to maturity at 18 years of age. 

Children are exposed to PBDEs in generally the same way as are adults, probably mainly by 

eating contaminated food. Because of their smaller weight, children’s intake of PBDEs per 

kilogram (or pound) of body weight may be greater than that of adults. The most likely way that 

infants might be exposed to PBDEs is from breast milk containing lower brominated congeners, 

although fetuses in the womb could also be exposed. DecaBDE is poorly absorbed in the body 

and, therefore, is unlikely to be found in breast milk or the fetus to any significant extent. 

Children who live near hazardous waste sites might accidentally eat some PBDEs by putting 
PBBs and PBDEs 17 

2. PUBLIC HEALTH STATEMENT—PBDEs 

dirty hands or other soil/dirt covered objects in their mouths, or through eating without washing 

their hands. Some children also eat dirt on purpose. It is also possible that children could be 

exposed to PBDEs following transport of the chemical on clothing from the parent’s workplace 

to the home. 

As indicated above, children can be exposed to PBDEs—mainly the lower brominated 

congeners—both before birth and from breast milk. The lower brominated PBDEs are much 

more likely than decaBDE to be stored in the mother’s body and released during pregnancy, 

cross the placenta, and entering fetal tissues. Because lower brominated PBDEs dissolve readily 

in fat, they can accumulate in breast milk fat and be transferred to babies and young children. 

Lower brominated PBDEs have been found in breast milk. In most cases, however, the benefits 

of breast-feeding outweigh any risks from exposure in mother’s milk. You should consult your 

health care provider if you have any concerns about PBDEs and breast feeding. The nervous 

system and thyroid are still developing in the fetus and child. That means the effects of lower 

brominated PBDEs on these systems might be more significant if exposure occurs during the 

periods before and soon after birth. That suggests that fetuses and children are more susceptible 

to PBDEs than adults. 

More information regarding children’s health and exposure to PBDEs can be found in Chapter 5 

(Section 5.7). 

2.7 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO PBDEs? 

If your doctor finds that you have been exposed to substantial amounts of PBDEs, ask whether 

your children might also have been exposed. Your doctor might need to ask your state health 

department to investigate. 

PBDEs are currently in widespread production and use. However, production of commercial 

pentaBDE and octaBDE products will soon begin to be phased out in the United States and 

elsewhere. You and your children may by exposed to lower brominated PBDEs by eating fish or 

wildlife from contaminated locations. Children who live near hazardous waste sites should be 
PBBs and PBDEs 18 

2. PUBLIC HEALTH STATEMENT—PBDEs 

discouraged from playing in the dirt near these sites because they could contain PBDEs. 

Children should also be discouraged from eating dirt, and careful handwashing practices should 

be followed. 

As mentioned in Section 2.3, workplace exposure to PBDEs can occur during the production of 

commercial PBDE mixtures and of PBDE-containing plastic products. Workers involved in 

recycling plastic products, or who repair computers in confined workplaces can also be exposed 

to PBDEs. If you are exposed to PBDEs while at work, you may carry them home on your 

clothes or body. Your occupational health and safety officer at work can tell you whether the 

products you work with may contain PBDEs and whether those are likely to be carried home. If 

this is the case, you should shower and change clothing before leaving work. Your work clothes 

should be kept separate from other clothes and laundered separately. 

2.8 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN 

EXPOSED TO PBDEs? 

Special tests can determine whether PBDEs are in the blood, body fat, and breast milk. These 

are not regular or routine clinical tests, but could be ordered by a doctor to detect PBDEs in 

people exposed to them in the environment and at work. If your PBDE levels are higher than the 

normal levels, this will show that you have been exposed to high levels of the chemicals. 

However, these measurements cannot determine the exact amount or type of PBDEs that you 

have been exposed to, or how long you have been exposed. Blood tests cannot distinguish 

between recent or past exposures to PBDEs because these chemicals remain in the body a long 

time. Although tests can indicate whether you have been exposed to PBDEs to a greater extent 

than the general population, they do not predict whether you will be harmed. Blood tests are the 

easiest, safest, and probably the best method for detecting recent exposures to large amounts of 

PBDEs. Results of such tests should be reviewed and carefully interpreted by physicians with a 

background in environmental and occupational medicine. Nearly everyone has been exposed to 

pentaBDE commercial mixtures because they are found throughout the environment. That 

means people are more likely to have detectable amounts of the lower brominated PBDEs in 

their blood, fat, and breast milk. Recent studies have shown that levels of lower brominated 
PBBs and PBDEs 19 

2. PUBLIC HEALTH STATEMENT—PBDEs 

PBDEs in the general population of the United States continue to rise. The U.S. levels are 10–

100 times higher than levels in individuals living in Europe. 

More information on tests used to determine whether you have been exposed to PBDEs can be 

found in Chapter 5 (Section 5.11) and Chapter 9 (Section 9.1). 

2.9 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO 

PROTECT HUMAN HEALTH? 

The federal government develops regulations and recommendations to protect public health. 

Regulations can be enforced by law. The EPA, the Occupational Safety and Health 

Administration (OSHA), and the Food and Drug Administration (FDA) are some federal 

agencies that develop regulations for toxic substances. Recommendations provide valuable 

guidelines to protect public health, but cannot be enforced by law. The Agency for Toxic 

Substances and Disease Registry (ATSDR) and the National Institute for Occupational Safety 

and Health (NIOSH) are two federal organizations that develop recommendations for toxic 

substances. 

Regulations and recommendations can be expressed as “not-to-exceed” levels, that is, levels of a 

toxic substance in air, water, soil, or food that do not exceed a critical value that is usually based 

on levels that affect animals. Those levels are then adjusted to levels that will help protect 

humans. Sometimes these not-to-exceed levels differ among federal organizations because they 

used different exposure times (an 8-hour workday or a 24-hour day), different animal studies, or 

other factors. 

Recommendations and regulations are also updated periodically as more information becomes 

available. For the most current information, check with the federal agency or organization that 

provides it. Some regulations and recommendations for PBDEs include the following: 

EPA requires that companies that transport, store, or dispose of p-bromodiphenyl ether (a 

particular PBDE compound not found in any commercial PBDE product) follow the rules and 
PBBs and PBDEs 20 

2. PUBLIC HEALTH STATEMENT—PBDEs 

regulations of the federal hazardous waste management program. EPA also limits the amount of 

p-bromodiphenyl ether put into publicly owned waste water treatment plants. To minimize 

exposure of people to p-bromodiphenyl ether, EPA requires that industry tell the National 

Response Center each time 100 pounds or more of p-bromodiphenyl ether have been released to 

the environment. 

For more information on federal and state regulations and guidelines for PBDEs, see Chapter 10. 

2.10 WHERE CAN I GET MORE INFORMATION? 

If you have more questions or concerns about PBDEs, please contact your community or state 

health or environmental quality department, or contact ATSDR at the address and phone number 

below. 

ATSDR can also tell you the location of occupational and environmental health clinics. These 

clinics specialize in recognizing, evaluating, and treating illnesses that result from exposure to 

hazardous substances. 

Toxicological profiles are also available on-line at www.atsdr.cdc.gov/toxpro2.html and on 

CD-ROM. You may request a copy of the ATSDR ToxProfilesTM CD-ROM by calling the tollfree information and technical assistance number at 1-888-42ATSDR (1-888-422-8737), by 

e-mail at atsdric@cdc.gov, or by writing to: 

 Agency for Toxic Substances and Disease Registry 

 Division of Toxicology 

 1600 Clifton Road NE 

 Mailstop F-32 

 Atlanta, GA 30333 

 Fax: 1-770-488-4178 
PBBs and PBDEs 21 

2. PUBLIC HEALTH STATEMENT—PBDEs 

Organizations for-profit may request copies of final Toxicological Profiles from the following: 

 National Technical Information Service (NTIS) 

 5285 Port Royal Road 

 Springfield, VA 22161 

 Phone: 1-800-553-6847 or 1-703-605-6000 

 Web site: http://www.ntis.gov/ 

PBBs and PBDEs 23 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

3.1 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO PBBs IN THE UNITED 

STATES 

Polybrominated biphenyls (PBBs) are brominated organic compounds used as flame retardant additives in 

plastics, textiles, and other materials. As additives, they are physically mixed into product applications, 

rather than chemically bound. Therefore, they have the potential to migrate from the plastic matrix into 

the environment when conditions are ideal. Commercial production of PBBs began in approximately 

1970 and manufacture was discontinued in the United States in 1976, subsequent to a major agricultural 

contamination episode that occurred in Michigan in 1973. Concern regarding the health effects of PBBs 

is mainly related to exposures that have resulted from the regionally localized Michigan episode. 

PBBs are classes of structurally similar brominated hydrocarbons in which 2–10 bromine atoms are 

attached to the molecular structure (i.e., biphenyl). Monobrominated structures (i.e., one bromine atom 

attached to the molecule) are often included when describing PBBs. There are 209 different molecular 

combinations, or congeners, that are possible for PBBs. However, unlike polychlorinated biphenyls 

(PCBs), only a subset of these 209 congeners exists in commercial mixtures. Based on the number of 

bromine substituents, there are 10 homologous groups of PBB congeners (monobrominated through 

decabrominated), with each homologous group containing one or more isomers. The mono-, di-, tri-, 

tetra-, penta-, hexa-, hepta-, octa-, nona-, and decabromo- congeners can exist in 3, 12, 24, 42, 46, 42, 24, 

12, 3, and 1 isomers, respectively. The general chemical structures of PBBs are similar when viewed in 

one dimension, as shown below, where m+n = 1–10: 

1'

2' 3'

4'

6' 5'

1

3 2

4

5 6 Br Brm n

where m + n = 1–10 

Due to the position and number of bromine atoms, there are important three-dimensional differences in 

the structures of PBBs that can influence the molecules’ receptor interactions and toxicological properties 

(see Section 5.5, Mechanisms of Action). 
PBBs and PBDEs 24 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

People are environmentally exposed to PBBs of different congeneric composition than the source 

commercial mixtures due to differential partitioning and transformation of the individual congeners in the 

environment, including transformation in food animals (e.g., dairy cattle in the case of PBBs). 

Additionally, as discussed in Section 5.4, because PBBs are lipophilic and some congeners are not readily 

metabolized, they are likely to be retained in the body for long periods of time (years). 

Three commercial PBB mixtures were manufactured: hexabromobiphenyl, octabromobiphenyl, and 

decabromobiphenyl. The two main commercial hexabromobiphenyl PBB mixtures had the trade names 

FireMaster BP-6 and FireMaster FF-1. FireMaster FF-1 was produced by grinding FireMaster BP-6 and 

adding 2% calcium polysilicate as an anticaking agent. The hexabromobiphenyl mixtures contained 

varying proportions of di- through octabrominated homologues. 2,2',4,4',5,5'-Hexabromobiphenyl is the 

most abundant congener in the mixtures (53.9–68.0%), followed by 7.0–27.3% of 2,2',3,4,4',5,5'-heptabromobiphenyl. Commercial octabromobiphenyl mixtures contained a large proportion (47.4–60.0%) of 

nonabromobiphenyl congeners, whereas commercial decabromobiphenyls contain predominately (96.8%) 

decabromobiphenyl congener. 

Limited data are available on health effects of commercial decabromobiphenyl and octabromobiphenyl 

mixtures, although the hexabromobiphenyl mixtures FireMaster BP-6 and FireMaster FF-1 have been 

extensively tested. Most of the information on human health effects of PBBs comes from studies of 

Michigan residents who accidentally ingested milk, meat, and eggs that came from farms that used animal 

feed contaminated with FireMaster FF-1. In 1973, livestock on certain farms in Michigan were exposed 

to FireMaster FF-1 after it was mistaken as a feed supplement and mixed with feed that was distributed 

within the state for several months before being discovered. Health problems in dairy cattle, reported in 

the fall of 1973, were the first signs that this episode occurred, but the accidental addition of PBBs to 

animal feed was not identified as the cause of the problem until the spring of 1974. 

Available information on the metabolism of PBBs in livestock is insufficient to ascertain whether the 

people affected in Michigan ingested the original PBB mixtures or metabolic products of the PBBs. 

Based on limited information in dairy cattle and additional data in laboratory animals as discussed in 

Section 5.4.2.2, it is reasonable to assume that mainly unchanged penta-, hexa-, and heptabromobiphenyl 

congeners were consumed in animal products during the contamination episode. PBBs were excreted in 

cattle manure and, as such, were also environmentally distributed in Michigan via waste disposal on 

farms. The general population outside of Michigan could possibly have been exposed to PBBs by the 
PBBs and PBDEs 25 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

oral route via the food chain, and the inhalation and dermal routes represent the most likely routes of 

exposure to PBBs in occupational settings. 

3.2 SUMMARY OF HEALTH EFFECTS 

Most of the information on human health effects of PBBs comes from studies of Michigan populations 

where PBB-contaminated cattle feed accidentally led to entrance into the feed/food chain, and some 

information is available on health effects in PPB-exposed chemical workers. Although the human studies 

consist largely of observations on groups that are not well defined and lack accurate intake data, they do 

provide a picture of the health status of the affected people and an indication of potential effects for the 

general population who may be exposed to lower levels of PBBs. Thus far, there is little convincing 

evidence linking exposure to PBBs and adverse health effects in Michigan farm residents. A variety of 

symptoms (e.g., neurological and neuropsychiatric, gastrointestinal, hepatic, dermal, and musculoskeletal) 

have been reported, but the prevalence of these symptoms has not been definitively linked to the extent or 

types of exposure. Other reported effects in humans include neurodevelopmental effects in exposed 

children and immunological changes. Physical examinations and laboratory tests have shown few 

abnormalities that corresponded to the complaints, and prevalences of symptoms have not been correlated 

with serum PBB levels. However, the possibility of long-term effects cannot be ruled out. 

Most toxicity studies of PBBs in animals involved oral exposure, and numerous effects have been 

documented including hepatic, renal, dermal/ocular, immunological, neurological, and developmental. 

Other effects of oral exposure to PBBs include decreased thyroid function, body weight loss, and liver 

cancer. Adverse hepatic effects, as well as dermal and ocular effects, also have been observed in a 

limited number of dermal studies in animals. No significant adverse effects were observed in animal 

inhalation studies of PBBs, but only two studies have been conducted, and the mixtures that were tested 

(octabromobiphenyl and decabromobiphenyl) are not the lower brominated products (i.e., Firemaster 

mixtures) expected to be the most toxic based on oral data. A number of PBB effects are dioxin-like and 

consistent with the Ah receptor-mediated mechanism of action, including altered vitamin A homeostasis, 

thymic atrophy, dermal and ocular effects (e.g., chloracne and inflammation of eyelids), and body weight 

changes (wasting syndrome). The main health effects of PBBs are discussed in detail below. 

Thyroid Effects. The thyroid gland is an unequivocal target of PBBs in animals, and evidence in 

humans is suggestive of a similar relationship. Effects in workers exposed to unspecified PBBs and/or 

decabromobiphenyl included increased serum thyrotropin, low or borderline low serum thyroxine (T4), 
PBBs and PBDEs 26 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

and increased thyroid antimicrosomal antibody titers. A spectrum of effects has been observed in rats 

exposed for acute and intermediate durations, ranging from decreases in serum levels of serum T4 and 

serum triiodothyronine (T3) to histological and ultrastructural changes in the follicles. The preponderance 

of these studies tested FireMaster FF-1 or FireMaster BP-6 in rats, although chronic exposure to 

FireMaster FF-1 induced thyroid follicular hyperplasia in mice. Similar thyroid effects also occurred in 

offspring of treated rats and pigs. 

Thyroid effects were produced in rats in acute-duration studies at doses as low as 3 mg/kg/day (reduced 

serum levels of T4 hormone), but not at 1 mg/kg/day, and in rats in intermediate-duration studies at doses 

as low as 0.05 mg/kg/day (increased number and decreased size of follicles). No information is available 

on possible changes in thyroid hormones following chronic exposure. Histological examinations in 

chronic-duration studies found no thyroid alterations in rats at doses as high as 1.5 mg/kg/day (highest 

tested dose), although follicular cell hyperplasia was induced in mice at ≥1.3 mg/kg/day. The noobserved-adverse-effect level (NOAEL) of 1 mg/kg/day was used as the basis for an acute-duration 

minimal risk level (MRL) for oral exposure. The acute-duration lowest-observed-adverse-effect level 

(LOAEL) for hepatic effects is identical to the LOAEL for acute thyroid toxicity, but is a less appropriate 

basis for the MRL because organ functional implications are not as clear. The intermediate-duration 

LOAELs for thyroid and hepatic effects are also comparable to each other, but neither of these LOAELs 

is suitable for an intermediate MRL because reproductive and developmental toxicity occurred at a lower 

dosage. The thyroid LOAEL for chronic-duration exposure is unsuitable for deriving a chronic MRL 

because decreased survival occurred at the same dose (lower doses were not tested), and thyroid, liver, 

and other effects occurred at lower doses in intermediate-duration studies. 

Hepatic Effects. Histologically and ultrastructurally documented liver damage is a consistent and 

prominent finding among animals exposed to PBBs by the oral route, but studies of Michigan residents 

who were likely to have ingested PBB-contaminated food are inconclusive. The human studies do not 

demonstrate any clear association between abnormal liver-associated serum indices (serum glutamicoxaloacetic transaminase [SGOT], serum glutamic pyruvic transaminase [SGPT], lactic dehydrogenase 

[LDH], bilirubin) or liver enlargement and PBB exposure. No information is available on hepatic effects 

of PBBs in humans exposed by the inhalation or dermal routes. Although the available studies on liver 

effects in humans are largely inconclusive, the animal data, as summarized below, suggest that humans 

may also be affected. 
PBBs and PBDEs 27 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

Hepatic effects ranging from microsomal enzyme induction and liver enlargement to fatty changes and 

necrosis have been observed in rodents and other laboratory animal species exposed orally to FireMaster 

PBBs in acute-, intermediate-, and/or chronic-duration studies. Acute- and intermediate-duration oral 

data for octabromobiphenyl mixtures are only available for rats and suggest that hepatic histopathologic 

effects are milder than for FireMaster mixtures at similar doses. Similarly, intermediate-duration oral 

data for decabromobiphenyl mixture suggest that this PBB mixture is a less potent hepatotoxicant than an 

octabromobiphenyl mixture. No pathologic effects were reported in the liver of rats exposed to an 

octabromobiphenyl mixture in acute- and intermediate-duration inhalation studies, or in an intermediateduration study with a decabromobiphenyl mixture, but it is unclear if histology was evaluated following 

the intermediate-duration octabromobiphenyl mixture exposure. Acute dermal exposure to commercial 

mixtures of hexabromobiphenyl, but not octabromobiphenyl, has been reported to produce gross necrotic 

changes in the liver of rabbits. Hepatocyte enlargement and degenerative changes (vacuoles or necrosis) 

are the most sensitive adverse hepatic effects that have been observed in the acute-, intermediate-, and 

chronic-duration oral studies with FireMaster PBBs. The lowest hepatic LOAELs for acute and 

intermediate durations are identical or essentially the same as the LOAELs for thyroid effects. The acuteduration LOAEL for thyroid toxicity is used as the basis for an acute oral MRL, but neither hepatic nor 

thyroid LOAELs for intermediate-duration exposure are suitable for MRL derivation because 

reproductive and developmental toxicity occurred at a lower dosage. Hepatotoxicity occurred at the 

lowest dosage tested in chronic studies with rats and mice, and the hepatic LOAEL in mice also caused 

thyroid effects. Neither the rat nor the mouse LOAEL is a suitable basis for a chronic MRL, however, 

due to decreased survival at the same dosage and weight loss and developmental toxicity in monkeys at a 

lower chronic dosage. 

Altered vitamin A homeostasis, primarily manifested as decreased hepatic storage of vitamin A, is 

another established effect of PBBs in animals. Vitamin A is essential for normal growth and cell 

differentiation, particularly differentiation of epithelial cells, and some PBB-induced epithelial lesions 

resemble those produced by vitamin A deficiency. Because it is the primary storage site for vitamin A, 

the liver has a major role in retinol metabolism. Esterification of dietary vitamin A, hydrolysis of stored 

vitamin A, mobilization and release into the blood of vitamin A bound to retinol-binding protein, and 

much of the synthesis of retinol-binding protein occurs in the liver. 

Immunological and Lymphoreticular Effects. Altered lymphocyte transformation responses 

among populations exposed to PBB following the Michigan contamination episode have been reported by 

some investigators. Others have not been able to confirm these findings. However, it is clear that no 
PBBs and PBDEs 28 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

correlation can be established between altered immune parameters and PBB levels in serum. Some have 

suggested that PBBs associated with white blood cells is possibly the cause of the immunological 

dysfunction resulting from exposure to PBBs. This would imply that total PBB in plasma is not 

necessarily a good marker for immune dysfunction. Continuous examination of this cohort may resolve 

the controversy. 

Studies in animals, mostly intermediate-duration studies in rodents, indicate that a variety of 

immunological parameters such as spleen and thymus weights, antibody production, and lymphoproliferative responses can be affected by treatment with commercial PBB mixtures. The only chronic 

study found increased splenic hematopoiesis in mice, but no histological changes in the spleen, thymus, 

or lymph nodes of rats. It is apparent, however, that some of these effects are only seen at PBB levels that 

cause overt toxicity. Steroids are known to influence the immune response. Corticosterone levels were 

elevated in plasma of mice that were exposed to FireMaster BP-6 in the diet for 30 days, although the 

increase was not enough to be responsible for the observed immunological effect (reduced antibodymediated response to sheep red blood cells). Thymic atrophy and a reduction in lymphocyte markers 

were reported in cows treated with PBB doses of 67 mg/kg/day for ≤60 days. However, these results 

should be interpreted with caution since the animals approached death at this dose. Based on the data 

available, it is difficult to suggest any particular species as the most sensitive. This is because different 

studies usually examined different end points, using different exposure protocols. It is unclear whether 

morphological changes in the reticuloendothelial system are more sensitive indicators of altered immune 

status than are functional changes. Although the limited data on humans are largely inconclusive, PBBs 

have altered immune responses in a variety of animal species, which suggests that humans may also be 

affected. 

Neurological Effects. Data from studies on Michigan residents exposed to PBBs as a result of the 

1973 feed contamination episode and data from a limited number of animal studies both suggest that 

exposure to PBBs may cause subtle effects on neuropsychological performance and development in 

humans. Symptoms of neurological effects, including fatigue, weakness, and decrements in the capacity 

to perform physical or intellectual work, were reported frequently by groups of farm families and 

residents of Michigan who were likely to have consumed farm products (milk, meat, and eggs) 

contaminated with PBBs; however, associations between PBB levels in serum or fat and the frequency of 

subjectively reported neurological symptoms were not found in several studies. The administration of 

neuropsychological tests to orally-exposed Michigan residents has not revealed abnormalities or 

associations between test performance and PBB levels in serum or fat. Similarly, no association between 
PBBs and PBDEs 29 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

performance in neuropsychological tests and serum PBB levels was made in a study of a small number of 

chemical workers exposed to unspecified PBBs via inhalation and/or dermal contact. 

Examinations of a small number of children (19) believed to have been exposed in utero or in early 

infancy during the peak of the Michigan PBB-feed contamination episode have not found consistent or 

marked effects on neuropsychological development. One study found a statistically significant 

association between performance in neuropsychological development tests and PBB levels in adipose 

tissues when the children were ≈2.5–4 years old, but a later examination when the children were ≈4–

6 years old did not find such an association for the same tests. 

Studies with rats have shown that oral exposure to PBBs at dose levels of ≈10 mg/kg/day for intermediate 

durations (1–6 months) produced decreased motor activity and weakness of the hind limb, but not operant 

behavior deficits or histopathological alterations of brain or spinal nerve tissue. Performance deficits in 

tests of learning behavior were observed in the offspring of female rats and female mice treated with oral 

doses of PBBs at approximate daily dose levels ranging from 0.2 to 10 mg/kg during gestation and 

lactation. Effects on acquisition of forward locomotion, cliff avoidance, cage emergence, and open-field 

activity were found in offspring of rats that were exposed to 2 mg/kg/day from day 6 of gestation through 

day 24 postpartum and observed until postnatal day 60. 

Dermal and Ocular Effects. Dermal lesions characterized as acne have been observed in humans 

occupationally exposed to PBBs. Increased prevalences of skin disorder symptoms, including rashes, 

acne, darkening or thickening of the skin, discoloration or deformity of fingernails or toenails, peeling and 

scaling, erythema, and hair loss, were reported by Michigan residents who were likely to have ingested 

PBB contaminated food. There was no association between serum PBB levels and prevalence of 

symptoms in one study, but physical examinations in the other study confirmed a slightly increased 

incidence of alopecia. Polymer fibers containing octabromobiphenyl mixture caused no dermal effects 

when placed on covered human skin for 6 days. 

Acute-, intermediate-, and chronic-duration oral studies have found no histological alterations in the skin, 

pinnae, ear canals, or salivary glands of rats or mice exposed to FireMaster FF-1 or FireMaster BP-6. 

Alopecia, loss of eyelashes, generalized subcutaneous edema, dry scaly skin, and periorbital edema 

developed in three monkeys that were exposed to low doses of FireMaster FF-1 for several months; 

related histological findings included sebaceous gland atrophy and metaplasia and keratinization of hair 

follicles. Uncharacterized dermatosis was observed in similarly treated pigs. Hyperkeratosis of the 
PBBs and PBDEs 30 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

eyelids and metaplasia of the tarsal glands with keratin cysts developed in cows that ingested FireMaster 

BP-6 for up to 60 days. Acute dermal application of FireMaster FF-1 induced hyperkeratosis, dilation 

and keratinization of hair follicles, and partial atrophy of the sebaceous glands in rabbits, but effects of 

octabromobiphenyl and decabromobiphenyl mixtures were generally mild (slight erythema and edema). 

Octabromobiphenyl mixture was not a sensitizer when applied to guinea pig skin. Although somewhat 

limited, the animal and human data are generally consistent and indicate that although PBBs can cause 

local responses such as irritation by direct dermal contact, exposure does not need to occur by the dermal 

route to produce cutaneous effects. 

Unspecified signs of ocular irritation were observed in rats intermittently exposed to a high (5,000 mg/m3

) 

dust concentration of decabromobiphenyl mixture for 4 weeks, but severity was not reported, and 

recovery was not assessed. Octabromobiphenyl and decabromobiphenyl mixtures caused mild eye 

irritation in rabbits when applied as a dry solid. Histopathological changes have not been observed in the 

eyes of rats or mice exposed orally to FireMaster FF-1 or FireMaster BP-6 in studies of acute, 

intermediate, or chronic duration. Xerophthalmia (extreme dryness of the conjunctiva) was reported in 

rats fed FireMaster BP-6 in an intermediate-duration study. Based on effects in animals, direct exposure 

to PBBs is likely to be irritating to human eyes. 

Body Weight Effects. Animal studies provide strong evidence that oral exposure to FireMaster 

PBBs causes a wasting syndrome characterized by progressive decreased weight gain, with immediate 

moderate to severe body weight loss generally preceding death. Effects on body weight have been 

observed in single dose, intermediate- and chronic-duration oral studies with rats, mice, guinea pigs, 

mink, monkeys, and/or cows. Changes in body weight were also observed in rabbits following acute 

dermal exposure to commercial mixtures hexabromobiphenyl, but not octabromobiphenyl, suggesting that 

the syndrome is independent of exposure route and is a potential effect of PBBs in humans. 

Reproductive Effects. A limited amount of data is available regarding the reproductive effects of 

PBBs in humans. The distribution of sperm counts, sperm motility, and sperm morphology was 

investigated in a small number (50) of male Michigan residents who ingested food produced on PBBcontaminated farms or who worked in a PBB manufacturing company. The study found no evidence for 

PBB-related effects compared with a putatively unexposed control group. No relationship was found 

between serum PBB levels and frequency and duration of breast-feeding in a retrospective study of 

women exposed to PBBs during the Michigan episode. A study of fetal mortality rates in Michigan 

counties did not include data for fetal mortalities occurring during the first trimester of pregnancy. 
PBBs and PBDEs 31 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

Animal studies provide limited evidence that FireMaster FF-1 and FireMaster BP-6 PBBs cause adverse 

reproductive effects in a variety of species. Increased menstrual cycle duration and prolonged 

implantation bleeding were observed in female monkeys fed approximate daily dose levels of 

0.012 mg/kg for 7 months before breeding and during pregnancy. A corresponding decrease in serum 

levels of progesterone suggests that the reproductive effects in the monkeys are related to PBB-induced 

endocrine imbalance. This dosage (0.012 mg/kg/day) is the lowest tested in any intermediate-duration 

study and also caused fetal deaths in the monkeys after ≈1 year of exposure. Although the reproductive 

effects are less serious, concern for serious developmental toxicity following exposures of <1 year 

precludes deriving an MRL for intermediate-duration exposure. Implantation was completely blocked in 

40–67% of female rats treated with gavage dose levels ≥28.6 mg/kg on alternate days between gestation 

days 0 and 14. Alterations in male reproductive organs were observed at doses that caused death in male 

rats (necrosis of the epithelial lining of the ductus deferens after 100 mg/kg for 4–5 weeks) and in a 

monkey (hypoactive seminiferous tubules after ≈0.73 mg/kg/day for 25 weeks). No alterations in litter 

size or fertility were observed in a study of male and female minks fed ≤0.39 mg/kg/day for 6–7 months 

prior to breeding and during pregnancy or in the F1 or F2 generations of female parental rats fed as much 

as 5 mg/kg/day during the postimplantation phase of gestation and through weaning. 

Based on the observations of adverse effects on reproduction in animals exposed to PBBs, the possibility 

that PBBs may cause reproductive harm in humans cannot be refuted and suggests that exposure of 

women to PBBs prior to and during the early phases of pregnancy may be of particular concern. 

Developmental Effects. Consistent or marked abnormalities have not been found in examinations of 

the physical and neuropsychological development of children exposed in utero or in early infancy during 

the peak of the Michigan PBB episode. Likewise, a comparison of fetal mortality rates for Michigan 

counties with a high percentage of quarantined farms and those for Michigan counties with no 

quarantined farms did not clearly establish or refute the possibility that the Michigan PBB episode caused 

developmental problems in exposed people. 

Fetotoxic and developmental effects have been observed in studies of FireMaster FF-1 or FireMaster 

BP-6 in several species of laboratory animals. Embryolethal effects or increased mortality among nursing 

young were observed in rats after oral exposure during gestation and in monkeys after exposure before 

conception and during pregnancy. Because the dosage (0.012 mg/kg/day) causing these serious 

developmental effects in monkeys is the lowest tested in any chronic study of PBBs, it is not possible to 
PBBs and PBDEs 32 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

derive an MRL for chronic-duration exposure. Structural malformations in fetuses, including cleft palate, 

were also observed in rats and mice after exposure to these PBBs during gestation. Increased incidences 

of fetuses with extra ribs were reported in a study of rats orally exposed to commercial 

octabromobiphenyl mixture during gestation, but oral exposure to commercial decabromobiphenyl 

mixture was not embryotoxic, fetotoxic, or teratogenic in rats. Studies with FireMaster FF-1 and 

FireMaster BP-6 found that body weight gain was reduced in the offspring of rats and mice after exposure 

during gestation, in rat offspring after exposure during gestation and lactation, and in mink kits after 

parental exposure before and during pregnancy. Liver effects, including increased liver weight and 

hepatic cytochrome P-450 enzymic activity, hepatocyte enlargement, vacuolization, and other 

degenerative changes, were observed in the offspring of rats, mice, and/or swine fed FireMaster FF-1 or 

FireMaster BP-6 during gestation and/or lactation. Performance deficits in tests of operant behavior were 

observed in offspring of rats and mice after oral exposure to FireMaster FF-1 or FireMaster BP-6 during 

pregnancy and lactation. 

Although the available human data regarding developmental effects of PBBs are inconclusive, the results 

from animal studies strongly suggest that PBBs may cause mild to severe developmental effects in 

humans, including growth retardation, alteration of neuropsychological development, and structural 

malformations. 

Cancer. There is no epidemiological evidence of an association between exposure to PBBs and 

increased prevalence of cancer (all sites) in Michigan residents who were likely to have ingested PBBcontaminated food. These data are inconclusive due to a short latency period of 4 years. Suggestive 

relationships between increasing serum levels of PBBs and risks of breast cancer, digestive system 

cancer, and lymphoma (not otherwise specified) were found in case-control studies of Michigan PBB 

registry enrollees who were followed for approximately 20 years. 

Oral studies with rats and mice demonstrate that FireMaster FF-1 is an unequivocal hepatocarcinogen. 

Hepatocellular adenomas, carcinomas, and/or liver neoplastic nodules were induced in these species 

following single or repeated (intermediate- and chronic-duration) exposures. These types of liver 

neoplasms even developed in the offspring of rats administered a single gavage dose during gestation or 

offspring of mice treated by diet during the perinatal period (throughout gestation and lactation). Liver 

neoplasm incidences were much higher in rats and mice exposed for up to 2 years in the only chronic 

bioassay than in the other studies that involved shorter-duration exposures of up to 6 months followed by 

an observation period of up to 2 years. Based on findings in male rats and mice of both sexes in this 
PBBs and PBDEs 33 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

study, there is some evidence that combined perinatal and adult dietary exposure to FireMaster FF-1 

enhanced the susceptibility of hepatocellular neoplasms in animals receiving adult exposure. 

Tumors were not clearly or consistently observed in nonhepatic tissues of animals exposed to FireMaster 

FF-1. Induction of thyroid follicular cell adenoma was inconclusive in mice in both National Toxicology 

Program bioassays. Equivocal increases in incidences of mononuclear cell leukemia were observed in 

adult-only exposed rats in the NTP chronic study, and combined perinatal and adult exposure showed no 

significant increase. Combined analysis of the incidences of this leukemia in the adult-only, perinatal 

only, and combined perinatal and adult exposure groups, however, showed an apparent association 

between increasing incidences and dose, and incidences in some of the groups exceeded historical control 

ranges. Evidence is available that oral administration of FireMaster BP-6 promotes development of 

initiated tumors in rats (liver enzyme-altered foci assays) and hamsters (tracheal papilloma assay). 

Based on the results of the oral studies of FireMaster FF-1 in rats, there is sufficient evidence to conclude 

that PBBs are carcinogenic in animals and potentially carcinogenic in humans. PBBs, as a group, have 

been classified as possibly carcinogenic to humans by IARC (Group 2B). This classification is based on 

sufficient evidence for carcinogenicity to animals and inadequate evidence of carcinogenesis in humans. 

NTP concluded that PBBs are reasonably anticipated to be human carcinogens based on sufficient 

evidence of carcinogenicity in animals. The EPA has not classified the carcinogenicity of PBBs. 

Because there is insufficient information about which constituents of the PBB mixtures are carcinogenic 

and the congener profile to which people may be exposed environmentally is likely to be different from 

the original PBB source, it is assumed that PBB mixtures of any composition are potentially carcinogenic. 

This assumption has uncertainty since it cannot be verified with current knowledge, and because the 

mechanism of PBB carcinogenesis in rodents has not been definitively elucidated. 

3.3 MINIMAL RISK LEVELS 

A number of the toxic effects of PBBs, including immunotoxicity, inhibition of body weight gain, and 

hepatic changes, appear to be mediated by a common mechanism of toxic action that involves a specific 

cytosolic molecular receptor (Ah receptor) (see Chapter 5, Section 5.5.2, Mechanisms of Toxicity). 

Although numerous factors can influence the toxicity of PBBs; including differences in absorption, 

distribution, and retention among animal species, at the tissue level, the potency of some individual 

congeners is determined by the magnitude of the response that is initiated by its binding with the Ah 

receptor. The binding affinity, in turn, is determined by the substitution pattern of the congener, and 
PBBs and PBDEs 34 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

many of the most toxic congeners resemble the structural configuration of 2,3,7,8-TCDD, and therefore 

are dioxin-like in toxicity. However, congeners that exhibit Ah receptor-mediated responses constitute a 

fraction of the components in commercial PBB mixtures. Therefore, it is presumed that congeners that 

act by other mechanisms also contribute to the toxicity of PBB mixtures, and the toxicity of PBBs is 

commonly classified as either “dioxin-like” or “nondioxin-like.” The mechanism(s) of toxicity for 

nondioxin-like PBB congeners is less clearly elucidated, but also may involve receptors (e.g., the estrogen 

receptor), or the involvement of reactive intermediates (e.g., arene oxides) that can form potentially toxic 

covalently bound substrate-macromolecular adducts. 

People are environmentally exposed to PBB mixtures of different congeneric composition than the 

original commercial PBB products. Although the toxicity or potency of environmental PBB mixtures 

consequently may be greater or less than that of commercial mixtures, there are insufficient mixture 

toxicity data on which to directly base MRLs for environmental PBBs. Due to the likelihoods that 

(1) multiple mechanisms (Ah-receptor-dependent mechanisms, Ah-receptor independent mechanisms, or 

both) may be involved in health effects induced by PBBs, (2) different PBB congeners may produce 

effects by different mechanisms, and (3) humans are exposed to complex mixtures of interacting PBBs 

with differing biological activities, as well as to the lack of a suitable approach for quantitatively 

evaluating joint toxic action from concurrent exposures to PBBs and other structurally similarly 

compounds (e.g., PCBs, CDDs, CDFs, and PBDEs) in the environment, data from commercial PBB 

mixtures are used to develop MRLs for assessing health risks from environmental exposures to PBBs. 

Inhalation MRLs 

No MRLs have been derived for inhalation exposure to PBBs because human and animal data for all 

durations are either insufficient or lacking. Insufficiencies in the human inhalation data include mixedchemical and unquantified exposures. The animal inhalation database is limited by inadequately reported 

studies and lack of any information on the mixtures likely to be most toxic (i.e., FireMaster PBBs). 

Oral MRLs 

• An MRL of 0.01 mg/kg/day has been derived for acute oral exposure (14 days or less) to PBBs. 

The acute oral MRL was based on a NOAEL for decreased serum levels of thyroid T4 hormone identified 

in groups of 8–11 male rats that were treated with 0, 1, 3, or 6 mg/kg/day doses of an unspecified mixture 

of PBBs in lecithin liposomes by gavage for 10 days (Allen-Rowlands et al. 1981). The MRL was 
PBBs and PBDEs 35 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

estimated by dividing the NOAEL by an uncertainty factor of 100 (component factors of 10 for animal to 

human extrapolation and 10 for human variability). Levels of serum T4 were significantly (p<0.05) 

reduced at ≥3 mg/kg/day, indicating that the lowest dose (1 mg/kg/day) is the NOAEL. Despite the fact 

that an appropriate statistical test (t-test rather than an ANOVA with multiple comparison tests) was used 

to analyze the data, ATSDR is confident with the designation of the NOAEL and LOAEL values. The 

data in the manuscript are presented graphically, with the animal numbers presented as a range (8–

11 animals/group); thus, an ANOVA could not be performed from published report. However, using the 

graphical data, the change in plasma T4 levels in the 3 mg/kg/day groups is clearly on the order of 20–

30%, which represents a biologically significant change. As such, the identification of 3 mg/kg/day as a 

LOAEL, and 1 mg/kg/day as a NOAEL, is not contraindicated by the lack of appropriate statistical 

analysis. 

Decreased serum T4 is considered adverse due to unequivocal evidence from numerous studies that the 

thyroid is a target of PBBs with a spectrum of effects, including decreases in serum T3 and T4 hormone, 

thyroid enlargement, effects in the follicular cells (e.g., reduced size, hyperplasia with columnar 

appearance and papillary projections), and accumulation of colloid droplets (Akoso et al. 1982b; Byrne et 

al. 1987; NTP 1983; Gupta and Moore 1979; Kasza et al. 1978a; Norris et al. 1975a, 1975b; Sepkovic and 

Byrne 1984; Sleight et al. 1978). Additional information on the derivation of the acute-duration oral 

MRL for PBBs is provided in Appendix A. 

Intermediate- and chronic-duration oral MRLs were not derived because serious developmental and 

reproductive effects were observed in monkeys that had been exposed to PBBs for durations that spanned 

the intermediate and chronic categories at the lowest dose tested in the database. This dose 

(0.012 mg/kg/day) caused increased menstrual cycle duration and implantation bleeding after 6–7 months 

of exposure and fetal deaths (fetal abortion and stillbirth) after ≈1 year of exposure in monkeys, with 

surviving infants having decreased birth weight and decreased postnatal weight gain (Allen et al. 1978, 

1979; Lambrecht et al. 1978). Additionally, weight loss occurred in the maternal monkeys. The 

0.012 mg/kg/day serious LOAEL for developmental and reproductive effects is lower than the lowest less 

serious LOAELs for thyroid effects in rats (0.05 mg/kg/day) (Akoso et al. 1982b) and hepatic effects in 

guinea pigs (0.04 mg/kg/day) (Sleight and Sanger 1976). As the most sensitive effect seen following 

intermediate-duration oral exposure was a serious LOAEL, without an accompanying NOAEL, concern 

for serious developmental and reproductive toxicity following exposures of <1 year therefore precludes 

deriving an MRL for intermediate-duration exposure. Derivation of an MRL for chronic oral exposure is 

similarly precluded by the serious developmental effect (stillbirth) that occurred following exposures 
PBBs and PBDEs 36 

3. RELEVANCE TO PUBLIC HEALTH—PBBs 

exceeding 1 year in duration. The serious LOAEL in monkeys is lower than the lowest chronic dosages 

tested in other species (0.5 and 1.3 mg/kg/day in rats and mice, respectively) that caused decreased 

survival (NTP 1992). ATSDR’s classification of LOAELs into less serious and serious effects is 

discussed in the introduction to Section 5.2. 
PBBs and PBDEs 37 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

4.1 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO PBDEs IN THE UNITED 

STATES 

Polybrominated diphenyl ethers (PBDEs) are brominated organic compounds used as flame retardant 

additives in plastics, textiles, and other materials. As additives, they are physically mixed into product 

applications, rather than chemically bound. Therefore, they have the potential to migrate from the plastic 

matrix into the environment when conditions are ideal. Production of PBDEs began in the 1970s and has 

continued to the present. Concern for the possible health effects of PBDEs has heightened recently due to 

evidence that components of pentabromodiphenyl ether (pentaBDE) commercial mixtures are 

ubiquitously distributed at very low levels in the environment, biota, human tissues, and breast milk. 

Body-burden data indicate that the general population is exposed to low levels of primarily lower 

brominated (e.g., tetra- and penta-) BDEs. For example, the median concentrations of BDE 47, BDE 153, 

BDE 183, BDE 209, and total PBDEs (sum of four congeners) were 0.63, 0.35, 0.17, <1, and 2.2 ng/g 

lipid weight, respectively. DecaBDE was found at levels above the limit of quantification (1 pmol/g 

lipid). Environmental concentrations of lower brominated PBDEs (i.e., tetraBDE and pentaBDE) appear 

to be leveling off in Europe, but appear to be increasing in certain areas in Canada and the United States, 

although data are too sparse to make broad statements regarding trends. The concentration of total 

PBDEs in air ranges from 5.5 pg/m3

 in rural environments to 52 pg/m3 in urban air. Because PBDEs are 

hydrophobic in nature, this class of compounds has not been detected in water to any significant extent. 

Environmental concentrations of higher brominated commercial mixtures (e.g., decabromodiphenyl ether 

or decaBDE) are principally concentrated in soils and sediment near industrial point sources. 

PBDEs are classes of structurally similar brominated hydrocarbons in which 2–10 bromine atoms are 

attached to the molecular structure (i.e., diphenyl ether). Monobrominated structures (i.e., one bromine 

atom attached to the molecule) are often included when describing PBDEs. There are 209 different 

molecular combinations, or congeners, that are possible for PBDEs, although only a limited number exist 

in commercial mixtures. Based on the number of bromine substituents, there are 10 homologous groups 

of PBDE congeners (monobrominated through decabrominated), with each homologous group containing 

one or more isomers. The mono-, di-, tri-, tetra-, penta-, hexa-, hepta-, octa-, nona-, and decabromocongeners can exist in 3, 12, 24, 42, 46, 42, 24, 12, 3, and 1 isomers, respectively. The general chemical 

structure of PBDEs is shown below: 
PBBs and PBDEs 38 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

O

Brn Brm

1'

2'

3'

4'

5'

6'

1

2

3

4

5

6

where m + n = 1–10 

Due to the ether linkage and the position and number of bromine atoms, there are important threedimensional differences in the structures of PBDEs that can influence the molecules’ receptor interactions 

and toxicological properties as discussed in Section 5.5, Mechanisms of Action. In general, PBDEs are 

not expected to have the same array of three-dimensional conformations as either PBBs or 

polychlorinated biphenyls (PCBs). 

People are environmentally exposed to PBDEs of different congeneric composition than the source 

commercial mixtures, which have specific congeneric compositions. People are also environmentally 

exposed to lower brominated PBDEs (e.g., tetra- and penta- brominated congeners) due to differential 

partitioning and transformation of the individual congeners in the environment, including transformation 

in food animals. Additionally, as discussed in Section 5.4, because PBDEs are lipophilic and some 

congeners are not readily metabolized, they are likely to be retained in the body for long periods of time 

(years). 

Three commercial PBDE mixtures have been and continue to be produced: decaBDE, octaBDE, and 

pentaBDE. DecaBDE has accounted for more than 80% of world PBDE usage. The composition of 

commercial decaBDE is ≥97% decaBDE, with the remainder mainly nonabromodiphenyl ether 

(nonaBDE). Commercial octaBDE is a mixture of congeners ranging from hexabromodiphenyl ether 

(hexaBDE) to nonaBDE, and mixtures of pentaBDE are comprised of tetrabromodiphenyl ether 

(tetraBDE) to hexaBDE congeners. Congeners with less than four bromine atoms are generally not found 

in commercial PBDEs. The main isomers in the commercial pentaBDE product are 2,2’,4,4’,5-pentabromodiphenyl ether (i.e., 2,2',4,4',5-pentaBDE or BDE 99; see Section 6-1) and 2,2’,4,4’-tetrabromodiphenyl ether (i.e., 2,2',4,4'-tetraBDE or BDE 47; see Section 6-1). DecaBDE seems to be largely 

resistant to environmental degradation. However, there is very little information on the environmental 

degradation of decaBDEs and the data that exist are controversial. Octa- and pentaBDE commercial 

mixture components, that have a total of four or less bromines, likely undergo differential partitioning 
PBBs and PBDEs 39 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

(e.g., bioconcentration in aquatic organisms) and possibly some type of transformation process (e.g., 

photodegradation) as well. The octa- and pentaBDE mixture components have more readily yielded the 

predominance of lower-brominated tetra- and penta-congeners, particularly BDE 47, BDE 99, and 

2,2’,4,4’,6-pentaBDE (i.e., 2,2’,4,4’,6-pentaBDE or BDE 100; see Section 6.1), which have been detected 

in environmental and human tissue samples (including breast milk). Most studies indicate that levels of 

lower brominated BDEs in body fluids are a factor of 10–100 higher for individuals living in the United 

States compared to individuals living in other regions of the world (e.g., Europe). The main source of 

human exposure to lower brominated PBDE congeners may be from dietary intake, particularly from 

foods with high fat content (e.g., fatty fish). Lower-brominated tetra- and penta- congeners have been 

detected in air samples, including those taken from remote areas, indicating that inhalation may also be a 

potential exposure route for the general population. 

4.2 SUMMARY OF HEALTH EFFECTS 

Information is available on the potential health effects of all three classes of commercial PBDE products, 

i.e., decaBDE (also referred to as DBDPO), octaBDE, and pentaBDE. As subsequently discussed, the 

toxicity of decaBDE is generally much less pronounced than for octa- and pentaBDE commercial 

products following acute and repeated-dose exposures. This dissimilar toxicity is likely related to the 

preferential accumulation of lower brominated congeners in the body, due to their greater partitioning and 

retention in lipid-rich tissues and lower rates of metabolism and elimination relative to decaBDE. In 

particular, in comparison with the lower brominated mixtures, oral studies in rats found that decaBDE is 

minimally absorbed (0.3–2%), has a relatively short half-life (<24 hours), and is rapidly eliminated via 

fecal excretion (>99% in 72 hours). These toxicokinetic differences appear to be related to the number 

and location of bromines on the diphenyl oxide molecule, and correlate with environmental monitoring 

data indicating that decaBDE has low bioaccumulation potential. Three-dimensional differences in 

molecular structure might also contribute to the dissimilar toxicity between congeners (see discussions in 

Section 4.3 and Section 5.5.2). 

The preponderance of health effects data on PBDEs is from studies of orally exposed laboratory animals. 

Based on the information summarized below and detailed in Chapter 5 (Health Effects), the animal data 

indicate that decaBDE is much less likely than lower brominated PBDEs to cause health effects in 

humans, and that main targets of concern for lower brominated PBDEs in humans are the liver, thyroid, 

and neurobehavioral development. Intermediate- and chronic-duration oral studies in rats and mice found 

that penta- and octaBDE commercial mixtures caused effects mainly in the liver and thyroid, particularly 
PBBs and PBDEs 40 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

enlargement and histological alterations in both organs and changes in serum levels of thyroid hormones. 

Little information is available on potential neurotoxic effects of PBDEs. PBDEs have not been tested for 

neurotoxicity using comprehensive test batteries, and most studies used a single dose level of a single 

congener. Mild impairments in spontaneous motor behavior and learning and memory were found in 

mice that were exposed to single low doses of 2,2',4,4'-tetraBDE (BDE 47), 2,2',4,4',5-pentaBDE 

(BDE 99), 2,2',4,4',5,5’-hexaBDE (BDE 153), and/or 2,2',3,3’,4,4',5,5’,6,6’-decaBDE (BDE 209) during 

perinatal and/or early postnatal periods and tested later in life. Concern for neurodevelopmental toxicity 

of PBDEs is further raised by the documented effects of lower brominated commercial mixtures on 

thyroid hormone homeostasis, and critical involvement of thyroid hormones in central nervous system 

development.. Several acute-duration studies with commercial pentaBDE mixtures and the single 

congener, 2,2’,4,4’-tetraBDE (BDE 47), suggest that immune suppression might be another important 

health end point for lower brominated BDEs, although comprehensive immunological evaluations have 

not been performed on any congener or commercial mixture. Information on the reproductive toxicity of 

PBDEs is limited to a one-generation study of a low-purity decaBDE product (77.4% decaBDE, 21.8% 

nonaBDE, 0.8% octaBDE) in rats that found no exposure-related functional effects. Developmental 

toxicity studies have shown no evidence of teratogenicity of penta- and octaBDEs in rats and rabbits, 

although fetotoxic effects, including skeletal ossification variations at maternally toxic doses, have 

occurred. No fetotoxic or teratogenic effects were induced in rats exposed to high, but not maternally 

toxic, doses of commercial decaBDE. Information on the carcinogenicity of PBDEs is limited to one 

chronic oral study of decaBDE and one chronic oral study of lower-brominated PBDEs. The decaBDE 

study found that lifetime exposure to a high-purity commercial product decaBDE caused neoplastic 

changes in the liver, including neoplastic nodules in rats and hepatocellular adenomas and carcinomas in 

mice. Equivocal increases in thyroid gland follicular cell tumors were also observed in mice. The study 

of lower brominated PBDEs used a former commercial product containing 77.4% decaBDE, 21.8% 

nonaBDE, and 0.8% octaBDE. No neoplastic effects occurred in rats exposed to this mixture for 

approximately 2 years, but the power of this study to detect carcinogenic effects is limited by very low 

dose levels and testing of only one species. The weight of available genotoxicity evidence indicates that 

PBDEs are not genotoxic. Dermal application studies in rabbits showed that decaBDE, octaBDE, and 

pentaBDE were nonirritating to intact skin and caused some erythematous and edematous responses in 

abraded skin. DecaBDE was not a skin sensitizer in humans and octaBDE and pentaBDE were 

nonsensitizing in guinea pigs, indicating that the PBDEs did not cause delayed contact hypersensitivity. 

Thyroid Effects. Limited information is available on thyroid effects in PBDE-exposed humans. 

There are suggestive occupational data as shown by effects that included increased serum thyroid 
PBBs and PBDEs 41 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

stimulating hormone (TSH), low or borderline low serum T4, and increased thyroid antimicrosomal 

antibody titers in workers exposed to decaBDE and/or unspecified PBBs. Due to the mixed nature of the 

exposure, the possible effects cannot be solely attributed to either chemical. There was no clear 

association between plasma levels of BDE 47 and thyroid hormone levels (free and total T3 and T4, TSH, 

free testosterone, follicle-stimulating hormone, lutenizing hormone, and prolactin) in men who consumed 

varying amounts of fatty fish from the Baltic Sea. Based on evidence in animals as summarized below, 

the thyroid is particularly sensitive to lower brominated PBDEs and a possible target of toxicity in 

exposed humans. 

Thyroid effects, which mainly included reduced serum T4 hormone levels and follicular cell hyperplasia, 

were consistently observed in rats and mice orally exposed to lower brominated commercial PBDE 

mixtures. Accompanying changes in serum TSH levels were not found, and the depression of serum T4

seems to involve enhanced metabolic formation of hydroxylated metabolites of PBDEs, which bind with 

high affinity to thyroid transport proteins. Acute duration studies showed decreases in serum T4 in rats 

exposed to ≥10 mg/kg/day octaBDE or ≥30 mg/kg/day pentaBDE for 4 days and in rats and mice exposed 

to ≥18 mg/kg/day pentaBDE for 14 days. Effects observed in intermediate-duration studies included 

thyroid hyperplasia in rats exposed to ≥8 mg/kg/day octaBDE for 30 days and reduced serum T4 in rats 

exposed to ≥10 mg/kg/day pentaBDE for 90 days. Exposure to pentaBDE on gestation day 6 through 

postnatal day 21 caused serum T4 reductions at 30 mg/kg/day in maternal rats and ≥10 mg/kg/day in their 

fetuses and neonatal offspring. Intermediate-duration exposure to a 77% decaBDE/22% nonaBDE 

commercial mixture caused thyroid hyperplasia in rats at doses ≥80 mg/kg/day for 30 days. Chronic 

(103-week) exposure to a high-purity (≥97%) commercial decaBDE mixture did not induce thyroid 

histopathological changes in rats at ≤2,550 mg/kg/day, although follicular cell hyperplasia was increased 

in mice exposed to 2,240 mg/kg/day of the same commercial product. 

The extent that PBDEs affect circulating levels of T4 or T3 might vary with species and rats are often 

regarded as more sensitive than humans. The main basis for this opinion seems to be studies showing that 

PBDEs affect binding of thyroid hormones to transthyretin (TTR), the primary transport protein in rats. 

Because TTR is not the major transport protein in humans, the findings have been interpreted as evidence 

that humans will be less sensitive than rats to thryoid effects of PBDEs. As discussed in Section 5.5.3, 

the greater sensitivity of rats is thought to be related to a smaller and more rapid turnover of the hormone 

pool in the rat thyroid, and to a more rapid clearance of secreted hormone in the rat; the latter being, in 

part, related to the absence of thyroid binding globulin (TBG) in rats. Whereas TTR is the major thyroid 

hormone binding protein in rats, TBG is the main binding protein in humans and most other mammals. 
PBBs and PBDEs 42 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

However, although TTR is not the main transport protein in human serum, it is the principal protein 

involved in transport of T4 to the brain in both rats and man. Therefore, without specific evidence that 

rats are more sensitive to PBDEs than humans, and considering the importance of TTR for the transport 

of thyroid hormones in the human brain, it is misleading to assume that PBDEs are unlikely to affect 

thyroid function in humans, or that humans are less sensitive to these effects than rats. 

Neurological Effects. A limited amount of information is available on neurological effects of 

commercial PBDE mixtures. No clinical signs of neurotoxicity or neurohistopathology were observed in 

rats or mice exposed to commercial decaBDE in dietary doses as high as 16,000–19,000 mg/kg/day for 

14 days, 8,000–9,000 mg/kg/day for 13 weeks, or 2,550–7,780 mg/kg/day for 103 weeks. Although the 

high doses and extended exposure durations provided opportunities for the induction and/or development 

of clinical signs, the study is limited by lack of testing for subtle behavioral changes and neurodevelopmental effects. A commercial pentaBDE mixture was evaluated for several behavioral end points 

in offspring of rats that were perinatally exposed to 1–100 mg/kg/day by gavage on gestation day (GD) 

6 through postnatal day (PND) 21. Evaluation of the offspring as adults showed no alterations in motor 

or sensory development as assessed by motor activity, habituation, and auditory startle response, although 

suggestive decreases in fear conditioning were observed. 

Neurobehavioral effects of individual PBDE congeners were evaluated in mice that were exposed during 

perinatal and/or early postnatal periods to 2,2',4,4'-tetraBDE (BDE 47), 2,2',4,4',5-pentaBDE (BDE 99), 

2,2',4,4',5,5’-hexaBDE (BDE 153), or 2,2',3,3’,4,4',5,5’,6,6’-decaBDE (BDE 209). Most of these studies 

used similar single oral dose experimental designs and evaluated spontaneous motor behavior and swim 

maze performance at 2–6 months of age. The findings collectively indicate that the nervous system is a 

target of particular PBDE congeners during a defined critical phase of neonatal brain development, as 

shown by mild impairments in spontaneous motor behavior and learning and memory in older mice. One 

study used a different experimental design in which mice were exposed to BDE 99 from GD 6 to 

PND 21, and evaluated using a variety of somatic (body weight gain, hair growth, day of eyelid and ear 

opening, day of incisor eruption) and neurobehavioral (righting reflex, forelimb stick grasping reflex, 

forelimb placing reflexes, negative geotaxis, screen grasping and climbing, pole grasping, ultrasonic 

vocalizations, homing test) end points during PNDs 2–22, as well as spontaneous activity end points on 

PNDs 22–120. Findings were suggestive of delayed sensorimotor development and altered spontaneous 

behavior. Based on the limited available information (most of the studies were reported as abstracts), and 

considering the known effects of lower brominated PBDEs on thyroid hormone homeostasis and the 
PBBs and PBDEs 43 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

critical role that thyroid hormones play in development of the central nervous system, neurobehavioral 

development is a potential effect of concern for lower brominated PBDEs in humans. 

Hepatic Effects. The hepatotoxic potential of lower brominated PBDE mixtures is well-documented 

in animals by oral exposure. The spectrum of observed hepatic effects includes microsomal enzyme 

induction, liver enlargement, and degenerative histopathologic alterations that progress to tumors. 

Repeated dietary exposure to PBDEs typically caused liver enlargement with or without degenerative 

changes, and effects were generally dose-related in incidence and severity, more frequent and pronounced 

in males than females, and more severe with octaBDE and pentaBDE than decaBDE. For example, 

subchronic oral studies in rats showed that commercial pentaBDE mixtures were hepatotoxic at doses 

≤10 mg/kg/day. Increased liver weight and hepatocellular enlargement with vacuolation occurred in rats 

exposed to commercial pentaBDE doses as low as 2–9 mg/kg/day for 4–13 weeks. Increased incidences 

of degeneration and necrosis of individual hepatocytes were observed 24 weeks following exposure to 

≥2 mg/kg/day of commercial pentaBDE for 90 days in rats. In contrast, high purity commercial decaBDE 

caused no liver pathology in rats and mice at estimated doses as high as 2,000–8,000 and 2,375–

9,500 mg/kg/day, respectively. High purity commercial decaBDE caused liver effects only following 

lifetime exposure to doses that were still very high. Exposure to 94–97% decaBDE for 103 weeks caused 

liver thrombosis and degeneration in rats at 2,240 mg/kg/day, and centrilobular hypertrophy and 

granulomas in mice at ≥3,200 mg/kg/day. No studies are available on hepatic effects of PBDEs in 

humans. Based on the evidence in animals, lower brominated PBDEs are potentially hepatotoxic in 

humans. 

Immunological and Lymphoreticular Effects. Information regarding the immunosuppressive 

potential of PBDE mixtures is essentially limited to evidence from acute-duration oral studies of 

pentaBDE in animals. The plaque-forming splenic cell antibody response to injected sheep red blood 

cells was significantly reduced in mice exposed to 72 mg/kg/day pentaBDE for 14 days; single doses as 

high as 500 mg/kg had no effect. In the same study, exposure to up to 72 mg/kg/day had no effect on 

natural killer cell (NKC) activity. In vitro production of IgG immunoglobulin from pokeweed mitogenstimulated splenocytes was reduced in mice exposed to 36 mg/kg/day pentaBDE for 14 days. Other 

14-day studies in mice found no changes in natural killer cell activity to murine YAC-1 target cells at 

≤72 mg/kg/day or numbers of splenic and thymic lymphocyte subsets at ≤36 mg/kg/day, although 

18 mg/kg/day of the single congener BDE 47 caused significantly reduced numbers of total lymphocytes 

and CD4+, CD8+, and CD45R+ subtypes in spleen. Chronic ingestion of decaBDE caused splenic 

lesions (hematopoiesis, fibrosis, lymphoid hyperplasia) in rats exposed to ≥1,200 mg/kg/day for 
PBBs and PBDEs 44 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

103 weeks, indicating that it is considerably less immunotoxic than pentaBDE. No studies are available 

on immunological effects of PBDEs in humans. Due to the limited amount of data in animals, 

insufficient information is currently available to adequately characterize the human immunotoxic 

potential of PBDEs. 

Developmental Effects. Oral developmental toxicity studies of deca-, octa-, and pentaBDE have 

shown no evidence of teratogenicity in animals. Gestational exposure to a high (1,000 mg/kg/day) but 

maternally nontoxic dose of decaBDE was fetotoxic in rats as shown by subcutaneous edema and delayed 

skull bone ossification; however, it is of note that this preparation was only 77% decaBDE. A later study 

of decaBDE, using a >97% pure compound, reported no maternal or fetal toxicity at doses up to 

1,000 mg/kg/day. Commercial mixtures of octaBDE caused skeletal ossification variations in rats and 

rabbits at maternally toxic levels and other indications of fetotoxicity at lower doses. Effects of 

gestational exposure to octaBDE included minimally increased postimplantation loss in rats at 

≥10 mg/kg/day, increased resorptions in rats at 25 mg/kg/day, and increased skeletal variations in rabbits 

at 15 mg/kg/day and rats at 50 mg/kg/day. No evidence of fetotoxicity was found in the only available 

study of pentaBDE in rats at maternally toxic doses ≤200 mg/kg/day. No studies are available on 

developmental effects of PBDEs in humans. Based on the evidence in animals, PBDEs are unlikely to 

cause developmental toxicity at expected levels of exposure. 

Cancer. The only information regarding carcinogenicity of PBDEs in humans is available from a casecontrol study that found no clear association between risk of non-Hodgkin’s lymphoma and exposure to 

BDE 47 in a small group of Swedish men and women. 

For most PBDEs, including pentaBDE and octaBDE, animal studies of carcinogenic effects are not 

available; cancer data on PBDEs in animals are limited to results of studies on commercial decaBDE 

products. In a bioassay conducted by the NTP, male and female rats were exposed to high purity 

commercial decaBDE (lots that were 96 or 94–97% pure) in the diet in low doses of 1,120 and 

1,200 mg/kg/day, respectively, and high doses of 2,240 and 2,550 mg/kg/day, respectively, for 

103 weeks. Male and female mice were similarly exposed to low doses of 3,200 and 3,760 mg/kg/day, 

respectively, and high doses of 6,650 and 7,780 mg/kg/day, respectively. Incidences of neoplastic 

nodules in the liver were significantly increased in the male and female rats, although the term neoplastic 

nodule is poorly defined and understood, and is no longer used by NTP to characterize hepatoproliferative 

lesions in rats. Incidences of hepatocellular adenoma or carcinoma (combined) were significantly 

increased in the male mice. Slightly elevated incidences of thyroid gland follicular cell adenoma or 
PBBs and PBDEs 45 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

carcinoma (combined) were additionally observed in exposed male mice, although the increases were not 

statistically significant. Carcinogenicity was additionally evaluated in rats that were exposed to 0.01, 0.1, 

or 1.0 mg/kg/day dietary doses of a 77.4% decaBDE mixture (containing 21.8% nonaBDPO and 0.8% 

octaDBDPO) for approximately 2 years. No exposure-related neoplastic changes were found, but the 

power of this study to detect carcinogenic effects is limited by the very low dose levels in comparison to 

those tested in the NTP bioassay. 

Based on the limited evidence of carcinogenicity in animals in the NTP bioassay (significantly increased 

incidences of neoplastic liver nodules in rats and combined hepatocellular adenomas and carcinomas in 

mice), as well as the lack of human data, decaBDE has been classified in EPA Group C (possible human 

carcinogen) and IARC Group 3 (not classifiable as to its carcinogenicity to humans). EPA Group D 

classifications (not classifiable as to human carcinogenicity) were assigned to nona-, octa-, hexa-, penta-, 

tetra-, tri-, and p,p’-diBDE based on no human data and no or inadequate animal data. The U.S. 

Department of Health and Human Services has not classified the carcinogenicity of any PBDE mixture. 

4.3 MINIMAL RISK LEVELS (MRLs) 

PBDEs share some toxicological properties with other structurally similar polyhalogenated aromatic 

compounds, particularly PBBs, PCBs, chlorinated dibenzo-p-dioxins (CDDs), and chlorinated 

dibenzofurans (CDFs) (Agency for Toxic Substances and Disease Registry 1994, 1998, 2000). However, 

although these chemicals are structurally similar in two dimensions, PBDEs (and PCDEs) differ from the 

other classes on a three-dimensional basis. In particular, the oxygen bridge of the ether linkage in the 

diphenyl oxide molecule increases the distance between the biphenyl rings. This apparently reduces 

steric interactions between ortho substituents on the adjacent rings, such that the presence of ortho

bromines is unlikely to present a barrier to rotation that would prevent the two aromatic rings from 

assuming a fully coplanar configuration (Chen et al. 2001; Hardy 2002b; Howie et al. 1990). In other 

words, the ether bridge makes PBDEs more noncoplanar in nature, and introducing ortho substitutions 

into PBDEs does not create a spatial impediment for the two phenyl rings to assume a semi-flat position 

with respect to each other, as it does for PBBs or PCBs. Therefore, for PBDEs, the influences of the ether 

bridge and bromine position preclude clearly classifying the congeners as either dioxin-like (coplanar) or 

nondioxin-like (noncoplanar). This has implications not only for dioxin-type toxicities, which are 

mediated by the Ah (aryl hydrocarbon) receptor (AhR) pathway (and therefore might not be a significant 

issue for PBDEs), but also for nondioxin-type effects. The assumption that PBDEs share many 

toxicological characteristics with PCBs also does not consider geometrical differences due to the higher 
PBBs and PBDEs 46 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

atomic weight and considerably larger molecular volume of bromine compared to chlorine (Hardy 2000, 

2002b). These differences contribute to dissimilar physical/chemical properties that can influence the 

relative toxicokinetics and toxicities of the chemicals. The geometrical differences in bromine and 

chlorine also have implications for understanding the mechanism(s) of effects for nondioxin-like PBB 

congeners, which are not as well characterized as for PCBs. In particular, it cannot necessarily be 

assumed, on the basis of two-dimensional structure, that the mechanisms and effects for nondioxin-like 

PBBs and PCBs are similar. 

People are environmentally exposed to PBDE mixtures of different congeneric composition than the 

original commercial PBDE products. Although the toxicity or potency of environmental mixtures of 

congeners consequently might be greater or less than that of the original commercial PBDE product, there 

are insufficient mixture toxicity data on which to directly base MRLs for environmental PBDEs. Due to 

the likelihoods that (1) multiple mechanisms (Ah-receptor-dependent mechanisms, Ah-receptor 

independent mechanisms, or both) may be involved in health effects induced by PBDEs, (2) different 

PBDE congeners may produce effects by different mechanisms, and (3) humans are exposed to complex 

mixtures of interacting PBDEs with differing biological activities, as well as to the lack of a suitable 

approach for quantitatively evaluating joint toxic action from concurrent exposures to PBDEs, PBBs, 

PCBs, CDDs, and/or CDFs in the environment, data from commercial PBDE mixtures are used to 

develop MRLs for assessing health risks from environmental exposures to PBDEs. 

There are important differences in the environmental chemistry and toxicity of decaBDE compared to 

lower brominated BDEs. DecaBDE seems to be largely resistant to environmental degradation, whereas 

octa- and pentaBDE commercial mixture components are likely to undergo differential partitioning and 

transformation, such that tetra- and pentaBDEs are the predominant congeneric forms that have been 

detected in the environment. Tetra- and pentaBDE congeners are also the main PBDEs in human tissues 

(blood, adipose, and breast milk), and BDE 47, BDE 99, and BDE 100 are particularly prevalent in both 

environmental and biological samples. The preferential accumulation of the lower brominated BDEs is 

likely due to their partitioning and retention in lipid-rich tissues; decaDBE does not appreciably partition 

into lipid-rich tissues and has higher rates of metabolism and elimination. These characteristics seem to 

be a function of both the number and location of bromines on the diphenyl oxide molecule. The tetra- and 

pentaBDEs appear to be relatively well absorbed, whereas the fully brominated decaBDE, a large poorly 

soluble molecule, is very poorly absorbed (≈1% or less of an oral dose) and rapidly eliminated (≈99% of 

the dose within 72 hours). Further, decaBDE is significantly less toxic than lower brominated BDE 

mixtures. For example, subchronic oral studies in rats showed that commercial octa- or pentaBDE 
PBBs and PBDEs 47 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

mixtures were hepatotoxic at doses as low as 2–10 mg/kg/day, while high purity commercial decaBDE 

caused no liver pathology at doses as high as 8,000–9,500 mg/kg/day (IRDC 1976, 1977; NTP 1986; WIL 

Research Laboratories 1984). Liver effects (thrombosis and degeneration) of commercial decaBDE in 

rats were only induced following lifetime exposure, and doses were still very high (2,240 mg/kg/day) 

(NTP 1986). A study in weanling rats showed that acute oral exposure to commercial decaBDE had no 

effect on thyroid hormones, while similar exposures to commercial octaBDE or pentaBDE mixtures 

caused decreases in serum T4 and T3 levels (Zhou et al. 2001). Similarly, acute postnatal oral exposure to 

pure decaBDE congener (BDE 99) was less potent than BDE 47, BDE 99, and 2,2',4,4',5,5’-hexaBDE 

(BDE 153) in inducing behavioral effects in adult mice (Eriksson et al. 1999, 2001a, 2002a; Viberg et al. 

2001a, 2001b, 2002a, 2002b, 2003b). Due to the dissimilar toxicity and environmental chemistry of 

decaBDE as discussed above, separate MRLs were derived for decaBDE and lower brominated mixtures. 

Inhalation MRLs 

Decabromodiphenyl Ether 

No MRLs were derived for acute-, intermediate-, or chronic-duration inhalation exposure to decaBDE 

due to a lack of inhalation studies on this BDE congener. 

Lower Brominated BDEs 

Derivation of an acute-duration MRL for lower brominated BDEs is not recommended at this time due to 

insufficient information. The inhalation database for acute-duration exposure to PBDEs is essentially 

limited to two 14-day unpublished industry-sponsored studies of octaBDE in rats (Great Lakes Chemical 

Corporation 1978, 2000). Liver and nasal effect levels were identified in these studies, but are 

inappropriate bases for MRL estimation. As discussed below, derivation of an acute MRL is precluded 

by inconsistencies between the studies and a lack of suitable information on thyroid hormone levels or 

another sufficiently sensitive end point. 

In the Great Lakes Chemical Corporation (1978) study, groups of five male and five female Charles River 

CD rats were whole-body exposed to dust of an unspecified commercial octaBDE mixture in mean 

analytical concentrations of 0, 0.6, 3.7, 23.9, or 165.2 mg/m3 for 8 hours/day for 14 consecutive days. 

The average mass median aerodynamic diameter (MMAD) and geometric standard deviation (GSD) of 

the particles were 3.5 μm and 2, respectively. Study end points included clinical signs (including 

observations for respiratory distress and nasal and ocular irritation), body weight and food consumption, 
PBBs and PBDEs 48 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

hematology (5 indices), blood chemistry (5 indices, thyroid hormones not assessed), urinalysis 

(10 indices), organ weights (5 organs including thyroid/parathyroid), gross pathology, and histology 

(21 tissues including nasal turbinates, trachea, lungs, and thyroid). The clinical laboratory tests were 

limited to rats in the control and two highest dose groups. The histological exams were limited to the 

control and highest dose groups, except for the liver, which was examined in all groups. Signs of 

increased respiration rate (rapid breathing) were observed by the end of each exposure period in rats 

exposed to ≥24 mg/m3

; this effect always disappeared by the following morning. Liver weight was 

significantly increased and hepatic lesions occurred in rats exposed to ≥3.7 mg/m3

. At 3.7 mg/m3, the 

liver lesions consisted of very slight to slight, focal to multifocal cytoplasmic enlargement of the 

hepatocytes, accompanied by focal acidophilic degeneration of individual to small groups of cells. The 

liver lesions were similar in the higher dose groups except that that the hepatocyte enlargement was 

multifocal to diffuse in distribution, and the presence of focal, small to large areas of hepatocellular 

necrosis of very slight to marked degree. There were no exposure-related histological changes in other 

tissues. 

The Great Lakes Chemical Corporation (2000) study is currently only available as a secondary summary. 

This study was summarized in a recent European Union Risk Assessment Report (EU 2003a), but a report 

is not currently available in the Toxic Substances Control Act Test Submissions (TSCATS) database. 

The only pertinent information in the TSCATS database (as of March 2004) is a notice that the study was 

initiated (Great Lakes Chemical Corporation 2000), strongly suggesting that the completed study was 

never submitted. 

In the Great Lakes Chemical Corporation (2000) study, a commercial octaBDE product (bromine content 

78.7%, not otherwise identified) was administered to groups of five male and five female Crl:CD(SD)IGS 

BR rats, by nose-only inhalation as a dust aerosol, in measured concentrations of 0 (air only), 1.0, 10, 

110, or 250 mg/m3

 for 6 hours/day, 5 days/week, for 2 weeks. MMADs in the low to high level groups 

were 2.9, 3.2, 2.9, and 2.9 μm; GSDs were not reported in the available study summary. Clinical signs, 

body weight, food consumption, and survival were evaluated throughout the study. Hematology, serum 

chemistry, urine indices, and thyroid hormones were not assessed. Necropsies, organ weight 

measurements, and histological examinations were performed following exposure termination. The 

histological examinations included nasal cavity, larynx, trachea, lungs, liver, kidneys, adrenals, brain, 

spleen, testes, and ovaries, but excluded thyroid and heart. Exposure-related effects occurred in the liver 

and nasal cavity. Hepatic effects included increased mean absolute and/or relative liver weights at 

≥10 mg/m3

 in males and ≥110 mg/m3 in females, with the greatest increases at 110 and 250 mg/m3 (21–
PBBs and PBDEs 49 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

44%). Centribobular hypertrophy similarly occurred in the liver at ≥10 mg/m3 in both sexes (100% 

incidences in all groups except 4/5 females at 10 mg/m3

). Nasal effects included minimal to mild Goblet 

cell hyperplasia and/or hypertrophy at ≥1 mg/m3

 in males and ≥10 mg/m3 in females. The nasal Goblet 

cell hyperplasia/hypertrophy occurred in nasal levels II and III in males at 1 mg/m3

, and generally in nasal 

levels II–VI in both sexes at ≥10 mg/m3

. No changes were found in the remaining squamous, respiratory, 

and olfactory nasal epithelia, or other parts of the respiratory tract. 

As detailed above, hepatocellular hypertrophy was found at concentrations at ≥3.7 mg/m3

 in the Great 

Lakes Chemical Corporation (1978) study and ≥10 mg/m3

 in the Great Lakes Chemical Corporation 

(2000) study. This liver effect was accompanied by some degenerative hepatocellular changes in the 

1978 study, but this was not confirmed by the later study. Additionally, slight histological changes in 

nasal Goblet cells occurred at ≥1 mg/m3

 in the Great Lakes Chemical Corporation (2000) study, but there 

were no nasal effects in the earlier study. Interpretation of the significance and adversity of the liver and 

nasal alterations is complicated by the inconsistency of findings between the studies and other factors, 

including the small numbers of animals per study (five/sex/level) and different methods of inhalation 

exposure (whole-body versus nose-only). Additionally, a well designed 13-week study (Great Lakes 

Chemical Corporation 2001a, 2001b) found (1) hepatocellular hypertrophy, but no degenerative liver 

changes, at higher a minimum effect level (16 mg/m3

) than in both 14-day studies, and (2) no clearly 

exposure-related changes in nasal Goblet cells at concentrations below 202 mg/m3 (study summarized in 

intermediate MRL section). Further, the adversity of the slight nasal Goblet cell changes, which were 

predominantly minimal in severity and possibly suggestive of very slight nasal irritation in both the 

14-day and 13-week studies, is unclear. 

The available information indicates that there are insufficient bases for considering the hepatic and nasal 

changes as adverse acute effects. More importantly, exposure to ≥16 mg/m3

 caused changes in serum 

levels of thyroid hormones (decreased T3, increased TSH) in the 13-week study. Thyroid hormone levels 

were not determined in either of the 14-day studies. Therefore, due to the lack thyroid hormone data in 

the 14-day studies, as well as the lack of any clear lowest-observed-adverse-effect levels (LOAELs) for 

the other end points in the 14-day studies, particularly at exposures levels below the LOAEL for thyroid 

effects in the 13-week study, there is no sufficiently sensitive basis for derivation of an MRL for acuteduration exposure. 

• An MRL of 0.006 mg/m3

 has been derived for intermediate-duration inhalation exposure (15–

364 days) to lower brominated BDEs. 
PBBs and PBDEs 50 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

The intermediate-duration inhalation MRL is based on a no-observed-adverse-effect level (NOAEL) of 

1.1 mg/m3

 for changes in thyroid hormones in rats that were intermittently exposed to octaBDE for 

13 weeks (Great Lakes Chemical Corporation 2001a, 2001b). Calculation of the MRL is detailed below. 

The inhalation database for intermediate-duration exposure to PBDEs consists of one well-conducted 

13-week study (Great Lakes Chemical Corporation 2001a, 2001b). This is an unpublished industrysponsored study in which a commercial octaBDE product (bromine content 78.7%) was administered to 

groups of 10 male and 10 female Crl:CD(SD)IGS BR rats, via nose-only inhalation as a dust aerosol, in 

measured concentrations of 0 (air only), 1.1, 16, or 202 mg/m3

 for 6 hours/day, 5 days/week, for 

13 weeks. The mean MMADs in the low to high exposure groups were 2.0, 2.7, and 2.8 μm, and the 

corresponding mean GSDs were 3.37, 3.72, and 3.01. Clinical and physical signs, body weight, food 

consumption, and survival were evaluated throughout the study. Ophthalmic, hematology (11 indices), 

serum chemistry (18 indices), and serum thyroid hormone (TSH, total T3, and total T4) evaluations were 

performed near the end of the exposure period. Urinalyses were not conducted. Comprehensive 

necropies, organ weight measurements, and histological examinations (including respiratory tract and 

thyroid) were performed following exposure termination. 

Hepatic, nasal, lung, thyroid, and ovarian effects were observed (Great Lakes Chemical Corporation 

2001a, 2001b). The liver was affected in both sexes as shown by dose-related increases in centrilobular 

hepatocellular hypertrophy at ≥16 mg/m3

 and liver weight (absolute and relative) at 202 mg/m3. Total 

incidences of centrilobular hepatocellular hypertrophy in the 0, 1.1, 16, and 202 mg/m3

 groups were 1/10, 

0/10, 3/10, and 10/10 in males, and 0/10, 0/10, 3/10, and 6/10 in females; severity was predominantly 

minimal in all groups and was not dose-related. Changes in nasal Goblet cells were increased at 

202 mg/m3

, but showed no clear dose-related increasing trends for incidence or severity. Total incidences 

of nasal Goblet cell hypertrophy were slightly increased in nasal level II of both sexes at ≥1.1 mg/m3

; 

incidences in the 0, 1.1, 16, and 202 mg/m3

 exposure groups were 4/10 (all minimal), 9/10 (7 minimal, 

2 mild), 6/10 (all minimal), and 10/10 (9 minimal, 1 mild) in males, and 2/10 (all minimal), 6/10 (all 

minimal), 4/10 (all minimal), and 8/10 (all minimal) in females. Nasal Goblet cell hypertrophy was also 

slightly increased in nasal level IV in males at 202 mg/m3

 (4/10, 0/10, 1/10, and 8/10, all minimal 

severity, not increased in females). Histological changes in the lungs included alveolar histiocytosis and 

chronic active inflammation that were only clearly increased in incidence at 202 mg/m3

. Total incidences 

of alveolar histiocytosis at 0, 1.1, 16, and 202 mg/m3

 were 3/10, 5/10, 5/10, and 10/10 in males, and 0/10, 

5/10, 2/10, and 10/10 in females. Corresponding total incidences of chronic active lung inflammation 

were 0/10, 0/10, 2/10, and 10/10 in males, and 0/10, 1/10, 1/10, and 10/10 in females. The severity of 
PBBs and PBDEs 51 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

both lesion types tended to increase from minimal to mild/moderate at 202 mg/m3. Gross lung changes 

also occurred in both sexes at 202 mg/m3

; these included lung firmness and white discoloration and/or 

enlargement in the bronchial and/or mediastinal lymph nodes. The lymph node effects correlated with the 

histological finding of granulomatous inflammation. There were no exposure-related gross or 

histopathological changes in the spleen, bone marrow, thymus, or other tissues, including thyroid. 

Thyroid hormone assessments, however, showed exposure-related decreases in mean thyroxine (total T4) 

at ≥16 mg/m3

 in both sexes, and increases in TSH at ≥16 mg/m3 in males and 202 mg/m3 in females. The 

changes were usually statistically significant (p<0.05 or p<0.01) compared to controls and were 

considered to be consistent with chemical-induced hypothyroidism. There were no serum T3 changes. 

Qualitative histological evaluations of step sections of ovaries showed an absence of corpora lutea in 

3/10 females at 202 mg/m3

, compared to 0/10 in the control and lower exposure groups. This 30% 

incidence was interpreted to be a treatment-related effect because an absence of corpora lutea was 

considered unusual in rats at 20 weeks of age. 

Considering the questionable adversity of minimal severity nasal Goblet cell hypertrophy, lack of clear 

dose-related increasing trends for incidences and severity of this nasal effect, clear identification of both a 

NOAEL (1.1 mg/m3

) and LOAEL (16 mg/m3) for changes in serum levels of thyroid hormones, and 

abundant evidence for thyroid effects of PBDEs in oral studies, the effects on thyroid hormones are the 

most appropriate basis for estimation of an intermediate-duration inhalation MRL. The MRL of 

0.006 mg/m3

 was derived by dividing the NOAELHEC of 0.53 mg/m3 by an uncertainty factor of 30 (3 for 

species to species extrapolation with dosimetric adjustments and 10 for human variability) and a 

modifying factor of 3 (for an incomplete database reflecting a single study in one species). The 

NOAELHEC was calculated using the following equations: 

 NOAELADJ = 1.1 mg/m3 x 6 hours/24 hours x 5 days/7 days = 0.196 mg/m3

 NOAELHEC = NOAELADJ x RDDR = 0.196 mg/m3 x 2.7 = 0.53 mg/m3 

The RDDR for the extrathoracic (ET) region was used to extrapolate deposited doses in rats to deposited 

doses in humans. The following parameters were used to calculate the RDDR: MMAD of 2.0 μm with a 

mean GSD (sigma g) of 3.37, default human body weight of 70 kg, and a default female F344 rat body 

weight of 0.18 kg. Additional information on the derivation of the intermediate-duration inhalation MRL 

for lower brominated BDEs is provided in Appendix A. 
PBBs and PBDEs 52 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

No MRL was derived for chronic-duration inhalation exposure to lower brominated BDEs due to a lack of 

chronic studies. 

Oral MRLs 

Decabromodiphenyl Ether 

No MRL was derived for acute-duration oral exposure to decaBDE due to insufficient data. Information 

is available on effects of acute exposure on body and liver weights, microsomal enzyme induction in the 

liver, and serum thyroid levels in weanling rats (Carlson 1980b; NTP 1986; Zhou et al. 2001), but the 

database is limited by lack of LOAELs and/or sufficiently sensitive end points. 

• An MRL of 10 mg/kg/day has been derived for intermediate-duration oral exposure (15–

364 days) to decabromodiphenyl ether. 

The MRL was derived based on a NOAEL of 1,000 mg/kg/day for developmental toxicity in rats exposed 

to decaBDE for 19 days during gestation (Hardy et al. 2002). The MRL was estimated by dividing the 

NOAEL by an uncertainty factor of 100 (10 for animal to human extrapolation and 10 for human 

variability).. 

A commercial decaBDPE product (97.34% DBDPO, 2.66% nonaBDE and octaBDE) was administered to 

groups of 25 mated female Sprague-Dawley rats by gavage in corn oil in daily doses of 0, 100, 300, or 

1,000 mg/kg/day on gestation days 0 through 19 (Hardy et al. 2002). Each female was sacrificed on 

gestation day 20 and necropsied. End points examined included maternal clinical observations, maternal 

body weight/weight gain and food consumption, maternal gravid uterine and liver weights, maternal gross 

lesions, total number of corpora lutea, uterine implantations, early and late resorptions, viable and 

nonviable fetuses, and fetal weight and sex. Fetuses were examined grossly (all fetuses), evaluated for 

skeletal/cartilaginous malformations and ossification variations (approximately half of each litter), and 

evaluated for visceral malformations (remaining fetuses). No treatment-related effects on any maternal or 

fetal end points were observed, indicating that 1,000 mg/kg/day was the NOAEL for maternal and 

developmental toxicity. 

Only one intermediate-duration systemic toxicity study of high purity decaBDE has been conducted. In 

this study, a commercial decaBDE product (94B97% pure) was fed to F344 rats (10/sex/level) in 

estimated doses of 496–8,000 mg/kg/day, or B6C3F1 mice (10/sex/level) in estimated dietary doses of 
PBBs and PBDEs 53 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

589–9,500 mg/kg/day, for 13 weeks (NTP 1986). All animals were observed daily, weighed weekly, and 

necropsied at the end of the exposure period. Comprehensive histological examinations were performed, 

but limited to the control and high-dose groups. No hematology, clinical chemistry, or urine indices, or 

thyroid hormone levels, were evaluated. There were no compound-related clinical signs, deaths, body 

weight or food consumption changes, gross pathology, or histopathology, indicating that 8,000 and 

9,500 mg/kg/day are intermediate-duration NOAELs for systemic toxicity in rats and mice, respectively. 

As discussed above, the NOAEL for developmental toxicity is 1,000 mg/kg/day (Hardy et al. 2002). 

Because doses of decaBDE higher than 1,000 mg/kg/day have not been tested for developmental toxicity, 

and the NTP (1986) study indicates that this dose is also a NOAEL for systemic toxicity, the 

1,000 mg/kg/day developmental toxicity NOAEL is used as the basis for the MRL. Additional 

information on the derivation of the intermediate-duration oral MRL for decaBDE is provided in 

Appendix A. 

No MRL was derived for chronic-duration oral exposure to decaBDE. Only one chronic study of high 

purity decaBDE has been conducted. In this study, a commercial decaBDE product (94B97% pure) was 

fed to F344 rats (50/sex/dose level) in dietary doses of 1,120/1,200 mg/kg/day (males/females) or 

2,240/2,550 mg/kg/day, and B6C3F1 mice (50/sex/dose level) in doses of 3,200/3,760 mg/kg/day or 

6,650/7,780 mg/kg/day, for 103 weeks (NTP 1986). Animals were examined daily for clinical signs. 

Body weights and food consumption were measured throughout the study, and comprehensive gross and 

histological examinations were performed on all animals in all dose groups, including those that were 

moribund or died during the study. No hematology, clinical chemistry, or urine indices, or thyroid 

hormone levels, were evaluated. Liver degeneration and thrombosis were significantly (p<0.05) 

increased in male rats at 2,240 mg/kg/day; respective incidences in the control, low, and high dose groups 

were 13/50, 19/50, and 22/50 for degeneration, and 1/50, 0/50, and 9/50 for thrombosis. The thrombosis 

was characterized by a near total occlusion of a major hepatic blood vessel by a dense fibrin coagulum. 

Neoplastic nodules in the liver were increased in males at ≥1,120 mg/kg/day and in females at 

2,550 mg/kg/day in incidences that were dose-related and statistically significant, although no treatmentrelated increases in hepatocellular carcinomas were observed. Other effects in exposed rats included 

fibrosis of the spleen, lymphoid hyperplasia of the mandibular lymph nodes, and acanthosis of the 

forestomach at 2,240 mg/kg/day. In mice, histopathological changes occurred in 3,200 mg/kg/day in 

males in the liver (centrilobular hypertrophy and granulomas) and thyroid (follicular cell hyperplasia). 

An MRL was not derived because the lowest tested dose, 1,120 mg/kg/day in male rats, is a LOAEL for a 

liver lesion (neoplastic nodules) that is precancerous and associated with thrombosis in the same tissue. 
PBBs and PBDEs 54 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

Lower Brominated Diphenyl Ethers 

• An MRL of 0.03 mg/kg/day has been derived for acute-duration oral exposure (14 days or less) to 

lower brominated diphenyl ethers. 

The acute oral MRL is based on a NOAEL of 1 mg/kg/day for reduced serum levels of thyroid T4

hormone in fetal rats that were exposed to pentaBDE on days 4–20 of gestation (Zhou et al. 2002). The 

MRL was estimated by dividing the NOAEL by an uncertainty factor of 30 (component factors of 10 for 

animal to human extrapolation and 3 for human variability). A component factor of 10 was not used for 

human variability because the MRL is based on effects observed in a sensitive subgroup. Thyroid 

hormone levels were determined in Long-Evans rats that were administered a technical pentaBDE 

mixture (DE-71) in corn oil by gavage from GD 6 through PND 21, except for PND 0 (day of birth) 

(Zhou et al. 2002). Dams were sacrificed on GD 20 and PND 22, and offspring were sacrificed on GD 20 

and PNDs 4, 14, 36, and 90). Study end points included serum total T4 and T3 concentrations measured at 

each age point. The maternal exposure to pentaBDE caused reductions in serum total T4 that were 

significantly (p<0.05) different from controls in dams at 30 mg/kg/day on GD 20 and PND 22 (48 and 

44%, respectively, relative to controls), and in fetuses and offspring at ≥10 mg/kg/day on GD 20 (at least 

15% reduced) and PNDs 4 and 14 (50 and 64% maximal in the 10 and 30 mg/kg/day groups, 

respectively). The effect on T4 concentrations in the offspring was age-dependent as values returned to 

control levels by PND 36. There were no exposure-related effects on serum total T3 concentrations in the 

dams or offspring at any time, although T3 was not be measured in fetuses on GD 20 due to insufficient 

serum sample volume. The critical NOAEL of 1 mg/kg/day and LOAEL of 10 mg/kg/day for reduced 

serum T4 hormone levels in fetal rats that were exposed to pentaBDE (Zhou et al. 2002) are supported by 

a NOAEL of 3 mg/kg/day and a LOAEL of 10 mg/kg/day for reduced serum T4 levels in weanling 

(28-day-old) rats that were exposed to octaBDE for 4 days (Zhou et al. 2001). Data from other acuteduration studies of PBDEs that support the selection of the critical NOAEL and LOAEL include a 

NOAEL of 2.5 mg/kg/day and LOAEL of 10 mg/kg/day for fetotoxicity in rats exposed to octaBDE for 

10 days during gestation (Life Science Research Israel Ltd. 1987), a NOAEL of 5 mg/kg/day and LOAEL 

of 15 mg/kg/day for fetotoxicity in rabbits exposed to octaBDE for 13 days during gestation (Breslin et al. 

1989), and LOAELs of 18 mg/kg/day for reduced serum T4 in rats and mice exposed to pentaBDE for 

14 days (Fowles et al. 1994; Hallgren et al. 2001). Additional information on the derivation of the acuteduration oral MRL for lower brominated BDEs is provided in Appendix A. 

The human relevance of thyroid effects in rats is debatable. As discussed in the thyroid effects part of 

Section 4.2, mechanistic data indicate that humans are generally less sensitive than rodents because 
PBBs and PBDEs 55 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

(1) PBDEs likely affect thyroid function by displacing T4 from serum-binding proteins in blood and by 

increasing liver metabolism of thyroid hormones, (2) the major plasma thyroid hormone transporter 

protein in rats, transthyretin (TTR), is only a minor binding protein in humans, and (3) a thyroid hormone 

binding globulin, not TTR, is the major transporter protein in humans (Blay et al. 1993; Capen 1997; 

Sinjari et al. 1998). The importance of the role of TTR in human thyroid homeostasis is unclear because, 

although it is not is the main transport protein in humans, TTR is necessary for thyroid hormone transport 

to the developing human fetus. Consequently, it is appropriate to use reduced serum T4 in fetal rats as the 

critical effect for the MRL. Support for the relevance of this effect is provided by the lack of changes in 

thyroid hormones in the maternal animals at the dose that caused the effect on T4 in the offspring. 

• An MRL of 0.007 mg/kg/day has been derived for intermediate-duration oral exposure (15–

364 days) to lower brominated BDEs. 

The intermediate oral MRL is based on a LOAEL of 2 mg/kg/day for minimal liver effects in rats that 

were exposed to pentaBDE for 90 days (WIL Research Laboratories 1984). The MRL was estimated by 

dividing the LOAEL by an uncertainty factor of 300 (component factors of 3 for use of a minimal 

LOAEL, 10 for animal to human extrapolation, and 10 for human variability). Groups of 30 male and 

30 female Sprague-Dawley rats were exposed to pentaBDE (commercial mixture DE-71) in the diet at 

dosage levels of 0, 2, 10, or 100 mg/kg/day for up to 90 days. Study end points included clinical signs, 

body weight, food consumption, hematology, clinical chemistry (including serum T3 and T4), urine 

indices, gross pathology, and selected organ weights (brain, gonads, heart, liver, kidneys, thymus, and 

thyroid). Histological examinations were performed on the liver, thyroid, thymus, kidney, and lung in all 

dose groups and in all tissues (comprehensive evaluation) at 0 and 100 mg/kg/day. Hepatocytomegaly 

was observed in males at 2 mg/kg/day and in both sexes at ≥10 mg/kg/day. The hepatocytomegaly was 

similar in incidence and severity after 4 and 13 weeks of exposure, was dose-related with respect to 

severity (some affected hepatocytes had vacuoles that likely contained lipid), and was still observed in 

males at ≥10 mg/kg/day and females at 2 and 100 mg/kg/day at 24 weeks postexposure (in lessened 

severity and incidence). Examinations at 24 weeks following exposure also showed an increase in 

individual hepatocytes with degeneration and necrosis, effects that were considered to be likely exposurerelated and indicative of final loss of previously damaged cells. No NOAEL was identified because liver 

changes were observed at all dose levels. Nonhepatic effects included decreases in serum T4 levels at 

≥10 mg/kg/day and thyroid follicular cell hyperplasia at 100 mg/kg/day. Based on the observations of 

hypertrophy, mild degeneration, and slight necrosis, 2 mg/kg/day is considered to be a minimal LOAEL 

for liver effects. Data from other intermediate-duration studies that support selection of the 2 mg/kg/day 

minimal LOAEL include hepatic LOAELs of 5 mg/kg/day for cytomegaly (with vacuolation and necrosis 
PBBs and PBDEs 56 

4. RELEVANCE TO PUBLIC HEALTH—PBDEs 

at higher doses) in rats exposed to octaBDE for 13 weeks (IRDC 1977), and 9 mg/kg/day for 

hepatocellular enlargement and increased liver weight in rats exposed to octaBDE or pentaBDE for 

28 days (IRDC 1976). Other relevant effect levels from intermediate-duration studies include a thyroid 

LOAEL of 10 mg/kg/day for reduced serum T4 levels in fetuses and neonatal offspring of rats that were 

exposed to pentaBDE from GD 6 through PND 21 (Zhou et al. 2002). Additional information on the 

derivation of the intermediate-duration oral MRL for PBDEs is provided in Appendix A. 

A chronic-duration oral MRL was not derived for lower brominated BDEs due to insufficient data. Only 

one chronic study of PBDEs other than high purity decaBDE has been conducted (Kociba et al. 1975; 

Norris et al. 1975b). In this study, Sprague-Dawley rats (25/sex/dose level) were fed a 77.4% pure 

commercial decaBDE mixture (containing 21.8% nonaBDE and 0.8% octaBDE) for approximately 

2 years. Evaluations that included clinical signs, body weight, food consumption, hematology, clinical 

chemistry, urine indices, and comprehensive histological examinations showed no exposure-related 

effects. The highest NOAEL is 1 mg/kg/day (highest tested dose), but this NOAEL is not appropriate for 

MRL estimation due to insufficient sensitivity of the study. In particular, using the NOAEL of 

1 mg/kg/day and an uncertainty factor of 100, a chronic oral MRL based on this study would be 5 times 

higher than the 0.002 mg/kg/day intermediate MRL. A similar pattern was observed for thyroid effects in 

the study used to derive the acute-duration oral MRL (Zhou et al. 2001) as summarized above. Due to the 

insufficiencies of the chronic data for MRL derivation, the intermediate oral MRL could be used as a 

value for chronic exposure. 
PBBs and PBDEs 57 

5. HEALTH EFFECTS 

5.1 INTRODUCTION 

The primary purpose of this chapter is to provide public health officials, physicians, toxicologists, and 

other interested individuals and groups with an overall perspective on the toxicology of polybrominated 

biphenyls (PBBs) and polybrominated diphenyl ethers (PBDEs). It contains descriptions and evaluations 

of toxicological studies and epidemiological investigations and provides conclusions, where possible, on 

the relevance of toxicity and toxicokinetic data to public health. 

A glossary and list of acronyms, abbreviations, and symbols can be found at the end of this profile. 

PBBs and PBDEs are classes of brominated hydrocarbons that are used as flame retardant additives in 

plastics, textiles, and other materials. Both classes of chemicals are comprised of compounds in which 1–

10 bromine atoms are attached to the biphenyl structure in up to 209 different combinations. Based on 

the number of bromine substituents, there are 10 homologous groups of PBBs and PBDEs (monobrominated through decabrominated), each containing one or more isomers. PBBs and PBDEs are 

structurally similar when viewed in one dimension, differing only in the ether linkage between the two 

phenyl rings in PBDEs, but the oxygen bridge confers three-dimensional conformational differences that 

can influence toxicological properties. Consequently, on the basis of chemical structure, it cannot be 

assumed that the health effects of PBBs and PBDEs are necessarily similar. 

Commercial production of PBBs began in approximately 1970, and manufacture was discontinued in the 

United States in 1976 following a contamination episode that occurred in Michigan in 1973–1974. Three 

main commercial mixtures of PBBs were produced: hexabromobiphenyl, octabrombiphenyl, and decabromobiphenyl. The most prevalent hexabromobiphenyl PBB mixtures had the trade names FireMaster 

BP-6 and FireMaster FF-1. FireMaster FF-1 was produced by grinding FireMaster BP-6 and adding 2% 

calcium polysilicate as an anticaking agent. The hexabromobiphenyl mixtures contained varying 

proportions (depending on lot number) of di- through octabrominated homologues, and 2,2',4,4',5,5'-hexabromobiphenyl (BB 153) was the most abundant congener (53.9–68.0%) followed by 2,2',3,4,4',5,5'-

heptabromobiphenyl (7.0–27.3%). Commercial octabromobiphenyl PBB mixtures contained a large 

proportion (47.4–60.0%) of nonabromobiphenyl congeners, whereas commercial decabromobiphenyls 

contained predominately (96.8%) decabromobiphenyl congener. The general names hexabromobiphenyl, 
PBBs and PBDEs 58 

5. HEALTH EFFECTS 

octabromobiphenyl, and decabromobiphenyl are used in this profile to refer to unspecified commercial 

mixtures of these PBBs. 

Concern regarding the health effects of PBBs is largely related to exposures that resulted from the 

Michigan contamination episode. Livestock on farms in Michigan were exposed to FireMaster FF-1 over 

a period of approximately 10 months after it was accidentally mistaken for the feed supplement 

magnesium oxide and mixed with animal feed that was distributed within the state. Health problems in 

dairy cattle (decreased feed consumption and decreased milk production), reported in the fall of 1973, 

were the first signs that the contamination episode occurred, but accidental addition of PBBs to animal 

feed was not identified as the cause of the problem until late spring of 1974 (Fries 1985a; Jackson and 

Halbert 1974). The U.S. Food and Drug Administration (FDA) established tolerances of 1 ppm in milk 

and meat fat and 0.1 ppm in eggs in May 1974, which were revised downward to 0.3 and 0.05 ppm, 

respectively, in November 1974 due to improved analytical sensitivity (Dunckel 1975; Fries 1985a). The 

Michigan Department of Agriculture (MDA) subsequently lowered the FDA tolerance in meat fat from 

0.3 to 0.02 ppm, but there currently are no FDA or MDA tolerances for PBBs (FDA 1989; Fries 1985a). 

As a result of a farm animal testing and quarantining program established by the MDA in May 1974, 

about 30,000 dairy cattle, 2,000 swine, 400 sheep, and over 2,000,000 chickens were found to contain 

PBBs at concentrations requiring their destruction (Dunckel 1975; Fries 1985a; Mercer et al. 1976). 

Most of the information that is available on health effects of PBBs in humans comes from studies of 

Michigan residents who ingested milk, meat, and eggs that were produced on farms that used the 

FireMaster-contaminated animal feed. In the interval of more than 9 months between the accident, the 

detection and identification of its cause, and the beginning of testing and the establishment of quarantines, 

PBB-contaminated food products were consumed, not only by farm families and people that acquired 

produce directly from PBB-contaminated farms, but also by people who purchased food from markets 

(Anderson et al. 1979). The Michigan PBB contamination episode led to the establishment of 

epidemiological studies (that are still ongoing) of Michigan residents who were expected to have 

consumed PBB-contaminated food, as well as to a substantial increase in research activity regarding the 

health effects of PBBs in cattle, poultry, and laboratory animals. Compared to the FireMaster 

(commercial hexabromobiphenyl) PBBs, relatively limited data are available on health effects of 

commercial mixtures of octabromobiphenyl and decabromobiphenyl. Reviews of the research results on 

the toxicity of PBBs include those by Damstra et al. (1982), DiCarlo et al. (1978), Fries (1985a), Kay 

(1977), Kimbrough (1987), Kimbrough et al. (1978a), and WHO (1994a). 
PBBs and PBDEs 59 

5. HEALTH EFFECTS 

This profile discusses information regarding health effects of PBBs in humans and laboratory animals; 

some research on cattle and poultry is also discussed, but its relevance to human health effects is 

uncertain due to interspecies physiological differences. Although the toxicity in livestock from Michigan 

farms that used large amounts of contaminated feed is generally attributed to PBBs, data on effects in 

animals from farms with low PBB contamination have generated some controversy, because the signs of 

toxicosis in these animals have not been reproduced in cattle experimentally exposed to PBBs at levels 

that caused tissue residue concentrations ≈100 times greater than those in the farm animals (Jackson and 

Halbert 1974; Moorhead et al. 1977). This led some investigators to suggest that some signs of toxicosis 

reported in Michigan cattle reflected farm management procedures, nutritional deficiencies, microbial and 

parasitic infections, or exposure to unknown contaminants in the feed (Durst et al. 1977; Fries 1985a; 

Moorhead et al. 1977). Although exposure by ingestion occurred during the Michigan contamination 

episode, existing information on the metabolism of PBBs in livestock is insufficient to ascertain whether 

the people ingested PBBs or metabolic products of PBBs. However, based on available data discussed in 

Section 5.4.2.2, it is reasonable to assume that mainly unchanged penta-, hexa-, and heptabromobiphenyl 

congeners in animal products were consumed. 

Unlike PBBs, PBDEs have been continuously produced and used as flame-retardant additives since the 

1970s. Concern for health effects of PBDEs has heightened due to relatively recent evidence that some 

PBDE congeners have become ubiquitously distributed in the environment and are present in tissues and 

breast milk of the general population at levels that continue to increase. Three commercial PBDE 

mixtures have been produced: decabromodiphenyl ether (decaBDE), octabromodiphenyl ether 

(octaBDE), and pentabromodiphenyl ether (pentaBDE). DecaBDE has accounted for more than 80% of 

PBDE usage. The composition of commercial decaBDE is ≥97% of the pure congener with the 

remainder mainly nonaBDE. Commercial octaBDE is a mixture of congeners ranging from nona- to 

hexaBDE, and mixtures of pentaBDE are comprised of tetra-, penta-, and hexaBDE congeners. 

Congeners with less than four bromine atoms are generally not found in commercial PBDEs. Reviews on 

the health effects and other aspects of PBDEs include those by Darnerud et al. (2001), de Boer et al. 

(2000), de Wit (2002), Hardy (1999, 2002a, 2002b), McDonald (2002), NAS (2002), Rahman et al. 

(2001), Silberhorn et al. (1990), and WHO (1994b). Penta- and octa-BDEs are apparently being phased 

out of production and use. Great Lakes Chemical Corporation will cease production of penta- and 

octaBDE by the end of 2004 (Tullo 2003) and the European Union has banned the sale of all products 

containing more than 0.1% (by mass) penta- or octaBDE (EU 2003b). 
PBBs and PBDEs 60 

5. HEALTH EFFECTS 

Toxicity data for individual PBB and PBDE congeners are included in some discussions in this profile 

when these data corroborate or provide information on effects not documented for the PBB and PBDE 

mixtures. Congener-specific toxicity data are currently not practical for determining exposure levels of 

PBB and PBDE mixtures associated with adverse health effects at hazardous waste sites. This is due in 

part to the fact that standardized analytical procedures for congener mixtures and commercially available 

standards for all congeners are lacking, and congener-specific analyses are not routinely performed. 

Additionally, using current health effects evaluation procedures, toxicity data for individual congeners 

may overestimate or underestimate the actual health risk of PBB and PBDE mixtures because congeners 

vary in toxic potency and may be influenced by other congeners in an additive or less-than-additive way. 

It is also important to recognize that the PBBs and PBDEs to which people may be exposed may be 

different from the original PBB and PBDE source because of possible changes in congener composition 

resulting from differential partitioning and transformation in the environment and/or differential 

biological metabolism and retention. 

5.2 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE 

To help public health professionals and others address the needs of persons living or working near 

hazardous waste sites, the information in this section is organized first by route of exposure (inhalation, 

oral, and dermal) and then by health effect (death, systemic, immunological, neurological, reproductive, 

developmental, genotoxic, and carcinogenic effects). These data are discussed in terms of three exposure 

periods: acute (14 days or less), intermediate (15–364 days), and chronic (365 days or more). 

Levels of significant exposure for each route and duration are presented in tables and illustrated in 

figures. The points in the figures showing no-observed-adverse-effect levels (NOAELs) or lowestobserved-adverse-effect levels (LOAELs) reflect the actual doses (levels of exposure) used in the studies. 

LOAELs have been classified into "less serious" or "serious" effects. "Serious" effects are those that 

evoke failure in a biological system and can lead to morbidity or mortality (e.g., acute respiratory distress 

or death). "Less serious" effects are those that are not expected to cause significant dysfunction or death, 

or those whose significance to the organism is not entirely clear. ATSDR acknowledges that a 

considerable amount of judgment may be required in establishing whether an end point should be 

classified as a NOAEL, "less serious" LOAEL, or "serious" LOAEL, and that in some cases, there will be 

insufficient data to decide whether the effect is indicative of significant dysfunction. However, the 

Agency has established guidelines and policies that are used to classify these end points. ATSDR 

believes that there is sufficient merit in this approach to warrant an attempt at distinguishing between 
PBBs and PBDEs 61 

5. HEALTH EFFECTS 

"less serious" and "serious" effects. The distinction between "less serious" effects and "serious" effects is 

considered to be important because it helps the users of the profiles to identify levels of exposure at which 

major health effects start to appear. LOAELs or NOAELs should also help in determining whether or not 

the effects vary with dose and/or duration, and place into perspective the possible significance of these 

effects to human health. 

The significance of the exposure levels shown in the Levels of Significant Exposure (LSE) tables and 

figures may differ depending on the user's perspective. Public health officials and others concerned with 

appropriate actions to take at hazardous waste sites may want information on levels of exposure 

associated with more subtle effects in humans or animals (LOAELs) or exposure levels below which no 

adverse effects (NOAELs) have been observed. Estimates of levels posing minimal risk to humans 

(Minimal Risk Levels or MRLs) may be of interest to health professionals and citizens alike. 

Estimates of exposure levels posing minimal risk to humans (Minimal Risk Levels or MRLs) have been 

made for PBBs and PBDEs. An MRL is defined as an estimate of daily human exposure to a substance 

that is likely to be without an appreciable risk of adverse effects (noncarcinogenic) over a specified 

duration of exposure. MRLs are derived when reliable and sufficient data exist to identify the target 

organ(s) of effect or the most sensitive health effect(s) for a specific duration within a given route of 

exposure. MRLs are based on noncancerous health effects only and do not consider carcinogenic effects. 

MRLs can be derived for acute, intermediate, and chronic duration exposures for inhalation and oral 

routes. Appropriate methodology does not exist to develop MRLs for dermal exposure. 

Although methods have been established to derive these levels (Barnes and Dourson 1988; EPA 1990), 

uncertainties are associated with these techniques. Furthermore, ATSDR acknowledges additional 

uncertainties inherent in the application of the procedures to derive less than lifetime MRLs. As an 

example, acute inhalation MRLs may not be protective for health effects that are delayed in development 

or are acquired following repeated acute insults, such as hypersensitivity reactions, asthma, or chronic 

bronchitis. As these kinds of health effects data become available and methods to assess levels of 

significant human exposure improve, these MRLs will be revised. 

A User's Guide has been provided at the end of this profile (see Appendix B). This guide should aid in 

the interpretation of the tables and figures for Levels of Significant Exposure and the MRLs. 
PBBs and PBDEs 62 

5. HEALTH EFFECTS 

5.2.1 Inhalation Exposure 

A few studies have examined groups of chemical workers involved in the manufacture and distribution of 

PBBs and/or PBDEs (Bahn et al. 1980; Brown et al. 1981; Chanda et al. 1982; Landrigan et al. 1979; 

Rosenman et al. 1979; Stross et al. 1981). These people are believed to have been exposed predominately 

by dermal contact and inhalation, although the oral route cannot be ruled out. Results from these studies, 

therefore, are discussed in this section as well as in Section 5.2.3. The highest NOAEL and all LOAEL 

values from each reliable inhalation study of health effects end points in each species and duration 

category for PBDEs are recorded in Table 5-1 and plotted in Figure 5-1; due to the general lack of data 

regarding inhalation exposure to PBBs, no equivalent table is presented for PBBs. 

5.2.1.1 Death 

Polybrominated Biphenyls. No studies were located regarding death in humans after inhalation exposure 

to PBBs. 

Nose-only exposure to the highest attainable dust concentration of octabromobiphenyl mixture for 4 hours 

(960 mg/m3

 as a time-weighted average) was not lethal to six male rats observed for 7 days (Waritz et al. 

1977). No deaths occurred in groups of five male and five female rats that were exposed to a decabromobiphenyl dust mixture at concentrations ranging from 5 to 5,000 mg/m3

 concurrently with starch dust 

(6 mg/m3

) for 6 hours/day, 5 days/week for 4 weeks (Millischer et al. 1980). Information on lethality of 

inhaled hexabromobiphenyl PBB mixtures was not located. 

Polybrominated Diphenyl Ethers. No studies were located regarding death in humans after inhalation 

exposure to PBDEs. 

No deaths occurred in groups of five male and five female rats that were chamber-exposed to pentaBDE 

aerosol (compound dissolved in corn oil), octaBDE dust, or decaBDE dust in concentrations as high as 

200,000, 60,000, or 48,200 mg/m3

, respectively, for 1 hour and observed for the following 14 days (IRDC 

1974, 1975a, 1975b). Confidence in these studies is limited by a lack of control data. There was no 

mortality in rats that were exposed to dusts of commercial octaBDE products at levels of 174 mg/m3

 for 

8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978), ≤250 mg/m3

 for 

6 hours/day, 5 days/week for 14 days (Great Lakes Chemical Corporation 2000), or ≤202 mg/m3

 for 

6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2001a, 2001b). 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Inhalation

(mg/m3) (mg/m3) (mg/m3) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

ACUTE EXPOSURESystemic

1

3.7

Resp 3.7

2063

24

24 (reversible rapid breathing)

14 d8 h/d (CD)

Rat 

165

Cardio 165

165

Gastro 165

165

Hemato 165

0.6

Hepatic 0.6

3.7

3.7 (hepatocytomegaly and focal

hepatocellular degeneration)

165

Renal 165

165

Endocr 165

165

Ocular 165

165

Bd Wt 165
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Inhalation (continued)

(mg/m3) (mg/m3) (mg/m3) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

2

1

Resp 1

2119

14d5 d/wk6 h/d (CD)

Rat 

1

Hepatic 1

10

10

250

Renal 250

250

Bd Wt 250
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Inhalation (continued)

(mg/m3) (mg/m3) (mg/m3) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

INTERMEDIATE EXPOSURESystemic

3

16

Resp 16

2116

202

202 (alveolar histiocytosis, chronic

active lung inflammation)

13 wk5 d/wk6 h/d (CD)

Rat 

202

Cardio 202

202

Gastro 202

202

Hemato 202

202

Musc/skel 202

1.1

Hepatic 1.1

16

16 (centrilobular hepatocellular

hypertrophy)

202

Renal 202

1.1

1.1

b

Endocr

16

16 (decreased serum T4,

increased serum TSH)

202

Dermal 202

202

Ocular 202

202

Bd Wt 202

Immuno/ Lymphoret

4

16

16

2118

202

202 (grossly discolored and

enlarged bronchial andmediastinal lymph nodesassociated with chronic activelung inflammation and alveolarhistiocytosis)

13 wk5 d/wk6 h/d (CD)

Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Inhalation (continued)

(mg/m3) (mg/m3) (mg/m3) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Reproductive

5

16

16 F

a The number corresponds to entries in Figure 5 1.b Used to derive an intermediate-duration (15-364 days) inhalation minimal risk level (MRL) of 0.006 mg/m3 for lower brominated diphenyl ethers. The MRL wconverting the animal NOAEL of 1.1 mg/m3 to a duration-adjusted human equivalent concentration (NOAELHEC) of 0.53 mg/m3, and dividing by an uncertainspecies to species extrapolation with dosimetric adjustments and 10 for human variability) and a modifying factor of 3 (for an incomplete data base).Bd Wt = body weight; Cardio = cardiovascular; d = day(s); Endocr = endocrine; F = female; gastro = gastrointestinal; hemato = hematological; hr = hour(s); LOobserved adverse effect level; M = male; metab = metabolic; min = minute; mo = month(s); Musc/skel = muscular/skeletal; NOAEL = no observed adverse efferespiratory; T4 = thyroxine ; TSH = thyroid stimulating hormone; wk = week(s)

2117

202

202 F (absence of corpora lutea in

ovaries)

13 wk5 d/wk6 h/d (CD)

Rat 
0.1

1

10

100

1000

Respiratory

1r1r2r

Cardiovascular

1r

Gastrointestinal

1r

Hematological

1r

Hepatic

1r1r2r2r

Renal

1r2r

Endocrine

1r

Ocular

1r

mg/m3

Figure 5-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- Inhala

Acute (≤14 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic
0.0001

0.001

0.01

0.1

1

10

100

1000

Respiratory

3r3r

Cardiovascular

3r

Gastrointestinal

3r

Hematological

3r

Musculoskeletal

3r

Hepatic

3r3r

Renal

3r

Endocrine

3r3r

Dermal

3r

Ocular

3r

Body Weight

3r

mg/m3

Figure 5-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- Inhalation (Co

Intermediate (15-364 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic
PBBs and PBDEs 69 

5. HEALTH EFFECTS 

5.2.1.2 Systemic Effects 

Systemic effects that have been observed in humans and animals following inhalation exposure to PBBs 

and PBDEs are described below. 

Respiratory Effects. 

Polybrominated Biphenyls. No studies were located regarding respiratory effects in humans after 

inhalation exposure to PBBs. 

Slight dyspnea was observed in five male and five female rats that were exposed to a decabromobiphenyl 

dust mixture at 5,000 mg/m3

 concurrently with starch dust (6 mg/m3) for 6 hours/day, 5 days/week for 

4 weeks (Millischer et al. 1980). This effect was not observed at 500 mg/m3

 and lower concentrations or 

in air only-exposed controls, and there were no changes in pulmonary resistance and compliance in 

urethane-anesthetized rats, blood gases, and lung histology at any of the exposure levels. Lung function 

and blood gases were not evaluated in starch-exposed controls, but this is unlikely to be a serious study 

deficiency as the ratio of PBB to starch was ≈1,000 in the high exposure group. 

Polybrominated Diphenyl Ethers. No studies were located regarding respiratory effects in humans after 

inhalation exposure to PBDEs. 

Transient signs of respiratory distress that included tachynpea or dyspnea developed in rats that were 

chamber-exposed to pentaBDE aerosol (compound dissolved in corn oil), octaBDE dust, or decaBDE 

dust in very high concentrations of 200,000, 60,000, and 48,200 mg/m3

, respectively, for 1 hour (IRDC 

1974, 1975a, 1975b). Confidence in these effect levels is low due to a small number of tested animals 

and lack of control data. 

Two 14-day inhalation studies of commercial octaBDE have been conducted. In one study, rats were 

chamber-exposed to concentrations of 0, 0.6, 3.7, 23.9, or 165.2 mg/m3

 as powdered dust for 8 hours/day 

for 14 consecutive days (Great Lakes Chemical Corporation 1978). Increased respiration rate occurred at 

≥23.9 mg/m3

. The rapid breathing pattern developed by the end of each exposure period, always 

disappeared by the following morning, and was not observed at lower exposure levels. Histological 

examinations of the control and 165.2 mg/m3

 rats (other groups not examined) showed no changes in 
PBBs and PBDEs 70 

5. HEALTH EFFECTS 

tissues that included nasal turbinates, trachea, lungs, and mediastinal lymph nodes). In the other study, 

rats were nose-only exposed to 0, 1.0, 10, 110, or 250 mg/m3

 as dust aerosol for 6 hours/day, 5 days/week 

for 2 weeks (Great Lakes Chemical Corporation 2001a, 2001b). Histological examinations showed nasal 

alterations, consisting of minimal to mild goblet cell hyperplasia and/or hypertrophy at ≥1 mg/m3

 in males 

and ≥10 mg/m3

 in females. The nasal goblet cell changes occurred in nasal levels II and III at 1 mg/m3

and generally in levels II-VI at ≥10 mg/m3

, and were considered indicative of very slight nasal irritation. 

No histopathology in the lungs or toxicolgically significant clinical signs were observed. 

Histological changes in the lungs, but not clearly in the nasal cavity, were found in a study of rats that 

were nose-only exposed to 0, 1.1, 16, or 202 mg/m3

 as dust aerosol for 6 hours/day, 5 days/week for 

13 weeks (Great Lakes Chemical Corporation 2001a, 2001b). The pulmonary effects included alveolar 

histiocytosis and chronic active inflammation, occurred in both sexes, and were only clearly induced at 

202 mg/m3

. Total incidences of alveolar histiocytosis in the 0, 1.1, 16, and 202 mg/m3 exposure groups 

were 3/10, 5/10, 5/10, and 10/10 in males, respectively, and 0/10, 5/10, 2/10, and 10/10 in females, 

respectively. Respective total incidences of chronic active lung inflammation were 0/10, 0/10, 2/10, and 

10/10 in males, and 0/10, 1/10, 1/10, and 10/10 in females. Both lesions were predominantly minimal or 

mild in severity, with moderate severity occurring in a few high-dose animals. Additional effects 

included gross pulmonary changes in both sexes at 202 mg/m3

; these included lung firmness and white 

discoloration and/or enlargement in the bronchial and/or mediastinal lymph nodes. The gross lymph node 

changes correlated with the histological granulomatous inflammation. Effects in nasal tissues were 

equivocal. Incidences of nasal goblet cell hypertrophy were slightly increased in nasal level II of both 

sexes at ≥1.1 mg/m3

, but incidences were not clearly dose-related and there was essentially no increase in 

severity from minimal with increasing dose. Total incidences of goblet cell hypertrophy in nasal level II 

in the 0, 1.1, 16, and 202 mg/m3

 exposure groups were 4/10, 9/10, 6/10, and 10/10 respectively, in males, 

and 2/10, 6/10, 4/10, and 8/10, respectively, in females. Minimal severity goblet cell hypertrophy was 

also slightly increased in nasal level IV in males at 202 mg/m3

 (4/10, 0/10, 1/10, and 8/10), but not in 

females. 

Cardiovascular Effects. 

Polybrominated Biphenyls. No studies were located regarding cardiovascular effects in humans or 

animals after inhalation exposure to PBBs. 

Polybrominated Diphenyl Ethers. No studies were located regarding cardiovascular effects in humans 

after inhalation exposure to PBDEs. 
PBBs and PBDEs 71 

5. HEALTH EFFECTS 

No histopathological changes were observed in the heart of rats that were exposed to dusts of commercial 

octaBDE products at levels of 174 mg/m3

 for 8 hours/day for 14 consecutive days (Great Lakes Chemical 

Corporation 1978), or ≤202 mg/m3

 for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical 

Corporation 2001a, 2001b). 

Gastrointestinal Effects. 

Polybrominated Biphenyls. No studies were located regarding gastrointestinal effects in humans or 

animals after inhalation exposure to PBBs. 

Polybrominated Diphenyl Ethers. No studies were located regarding gastrointestinal effects in humans 

after inhalation exposure to PBDEs. 

No histopathological changes were observed in the stomach and lower gastrointestinal tract tissues of rats 

that were exposed to dusts of commercial octaBDE products at levels of 174 mg/m3

 for 8 hours/day for 

14 consecutive days (Great Lakes Chemical Corporation 1978), or ≤202 mg/m3

 for 6 hours/day, 

5 days/week for 13 weeks (Great Lakes Chemical Corporation 2001a, 2001b). 

Hematological Effects. 

Polybrominated Biphenyls. No studies were located regarding hematological effects in humans after 

inhalation exposure to PBBs. 

Hematology was normal in groups of five male and five female rats that were exposed to a decabromobiphenyl dust mixture at concentrations of 5 or 5,000 mg/m3

 concurrently with starch dust (6 mg/m3) for 

6 hours/day, 5 days/week for 4 weeks (Millischer et al. 1980). The evaluation included erythrocyte and 

leucocyte counts, differential leukocyte count, hematocrit, and hemoglobin level. 

Polybrominated Diphenyl Ethers. No studies were located regarding hematological effects in humans 

after inhalation exposure to PBDEs. 

No adverse hematological changes occurred in rats that were exposed to 24.4 or 174 mg/m3

 of 

commercial octaBDE dust aerosol for 8 hours/day for 14 consecutive days (Great Lakes Chemical 

Corporation 1978). Evaluation of a limited number of indices (hemoglobin, hematocrit, total erythrocyte 
PBBs and PBDEs 72 

5. HEALTH EFFECTS 

count, and total and differential leukocyte counts) showed no remarkable responses except for an 

elevation in leukocyte numbers. The observed increase in leukocyte counts was considered to be an 

unusual response by the investigators, although it was within the normal range for control rats in their 

laboratory. Comprehensive hematological assessments showed no unusual changes in rats exposed to 

commercial octaBDE as dust aerosol at concentrations of ≤202 mg/m3

 for 6 hours/day, 5 days/week for 

13 weeks (Great Lakes Chemical Corporation 2001a, 2001b). 

Hepatic Effects. 

Polybrominated Biphenyls. No studies were located regarding hepatic effects in humans after inhalation 

exposure to PBBs. 

No significant (p<0.05) increase in relative liver weight or hepatic histological changes were found in six 

male rats nose-only exposed to a octabromobiphenyl dust mixture at 960 mg/m3

 for 4 hours (timeweighted average, highest attainable concentration), and observed for 7 days (Waritz et al. 1977). 

Toxicity of octabromobiphenyl mixture vapor was investigated in groups of six rats almost continuously 

exposed (23 hours/day, 7 days/week) for 2, 4, 7, 9, 11, 13, or 15 weeks (Waritz et al. 1977). The 

exposure level was 0.00035 μg/m3

, which is the reported equilibrium concentration at 28 °C. Gross 

pathologic examination and measurement of relative liver weight showed no exposure-related changes at 

any of the sacrifices, but it is unclear if liver histology was evaluated. 

Relative liver weight was increased ≈25% in groups of 5 or 10 rats that were exposed to a decabromobiphenyl dust mixture at concentrations of 50–5,000 mg/m3

 concurrently with starch dust (6 mg/m3) for 

6 hours/day, 5 days/week for 4 weeks (Millischer et al. 1980). The increased liver weight was not 

accompanied by hepatic histologic changes, and therefore may be an adaptive response because PBBs are 

hepatic inducers and cause cellular proliferation (see Section 5.2.2.2 Hepatic Effects). No effects on liver 

weight or histology were observed at 5 mg/m3

. 

Polybrominated Diphenyl Ethers. No studies were located regarding hepatic effects in humans after 

inhalation exposure to PBDEs. 

Hepatic effects were observed in 14-day inhalation studies of dusts of commercial octaBDE mixtures. In 

one study, rats were chamber-exposed to concentrations of 0, 0.6, 3.7, 23.9, or 165.2 mg/m3

 as powdered 

dust for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978). Increased liver 

weight and hepatic histological changes occurred at 3.7 mg/m3

 and higher levels of exposure. At 
PBBs and PBDEs 73 

5. HEALTH EFFECTS 

3.7 mg/m3, the liver lesions consisted of very slight to slight severity focal to multifocal cytoplasmic 

enlargement of the hepatocytes, accompanied by focal acidophilic degeneration of individual to small 

groups of cells. The liver lesions were similar at ≥24.4 mg/m3

), except that the hepatocyte enlargement 

was multifocal to diffuse in distribution and accompanied by focal, small to large areas of hepatocellular 

necrosis of very slight to marked degree. In another study, rats were nose-only exposed to 0, 1.0, 10, 110, 

or 250 mg/m3

 as dust aerosol for 6 hours/day, 5 days/week for 2 weeks (Great Lakes Chemical 

Corporation 2001a, 2001b). Hepatic effects included increased mean absolute and/or relative liver 

weights at ≥10 mg/m3

 in males and ≥110 mg/m3 in females, with the greatest increases at 110 and 

250 mg/m3

 (21–44%). Centribobular hypertrophy similarly occurred in the liver at ≥10 mg/m3 in both 

sexes (100% incidences except for 4/5 females at 10 mg/m3

). 

Similar hepatic changes were found in a study of rats that were nose-only exposed to 0, 1.1, 16, or 

202 mg/m3

 commercial octaBDE as dust aerosol for 6 hours/day, 5 days/week for 13 weeks (Great Lakes 

Chemical Corporation 2001a, 2001b). The liver was affected in both sexes as shown by dose-related 

increased centrilobular hepatocellular hypertrophy at ≥16 mg/m3

 and increased liver weight (absolute and 

relative) at 202 mg/m3

. Respective total incidences of centrilobular hepatocellular hypertrophy 

(predominantly minimal to mild) in the 0, 1.1, 16, and 202 mg/m3

 groups were 1/10, 0/10, 3/10, and 

10/10 in males, and 0/10, 0/10, 3/10, and 6/10 in females. Serum chemistry evaluations showed no clear 

effects of exposure. Serum cholesterol was significantly increased (39.8% less than controls, p<0.01) in 

202 mg/m3

 females, but the magnitude of the elevation was not considered toxicologically significant. 

Some other statistically significant serum chemistry alterations (increased mean globulin and total 

protein, decreased albumin/globulin ratio) also occurred in the 202 mg/m3

 females, but were not 

considered exposure-related due to small magnitudes of changes and lack of similar changes in the males. 

Renal Effects. 

Polybrominated Biphenyls. No studies were located regarding renal effects in humans after inhalation 

exposure to PBBs. 

Groups of six rats were exposed to 0.00035 μg/m3

 of octabromobiphenyl mixture vapor (equilibrium 

concentration) 23 hours/day, 7 days/week for 2, 4, 7, 9, 11, 13, or 15 weeks (Waritz et al. 1977). Gross 

pathologic examination at each sacrifice and measurement of relative kidney weight at the last sacrifice 

showed no exposure-related changes, but it is unclear if kidney histology was evaluated. 
PBBs and PBDEs 74 

5. HEALTH EFFECTS 

Urinalysis was normal in groups of five male and five female rats that were exposed to a 

decabromobiphenyl dust mixture at concentrations ranging from 5 to 5,000 mg/m3

 concurrently with 

starch dust (6 mg/m3

) for 6 hours/day, 5 days/week for 4 weeks (Millischer et al. 1980). The analysis 

included pH, specific gravity, proteins, glucose, ketone bodies, biliary pigments, urobilinogen, blood, and 

microscopic examination of sediment. A comprehensive histology evaluation was performed in this 

study, but the only tissues specifically mentioned as having been examined are the liver and lungs. 

However, a total of 21 tissues were examined; therefore, it is probable that the kidney was examined, but 

was not discussed because no histological alterations were found. 

Polybrominated Diphenyl Ethers. No studies were located renal respiratory effects in humans after 

inhalation exposure to PBDEs. 

No histopathological changes were observed in the kidneys or urinary bladder of rats that were exposed to 

dusts of commercial octaBDE products at levels of 174 mg/m3

 for 8 hours/day for 14 consecutive days 

(Great Lakes Chemical Corporation 1978), ≤250 mg/m3

 for 6 hours/day, 5 days/week for 14 days (Great 

Lakes Chemical Corporation 2000), or ≤202 mg/m3

 for 6 hours/day, 5 days/week for 13 weeks (Great 

Lakes Chemical Corporation 2001a, 2001b). Urinalyses were not performed in any of these studies. 

Endocrine Effects. 

Polybrominated Biphenyls. Hypothyroidism was diagnosed in 4 of 35 men who were occupationally 

exposed to unspecified PBBs and/or decaBDE (Bahn et al. 1980). The cohort consisted of workers (mean 

age 35.9 years) who had been employed at a production plant for at least 6 weeks during a 52-month 

period during which PBBs and decaBDE were the only chemicals manufactured and who had volunteered 

for a comprehensive medical evaluation performed 3 months after the end of the 52-month period. There 

was no further description of exposure, and it was assumed to have occurred by both inhalation and 

dermal routes. The cohort was matched by sex, race, and age to 89 unexposed control subjects. Four 

subjects (22–50 years old, employed for 9–46 months not entirely during the 52-month production period) 

had elevated serum thyrotropin levels (mean 37.5 versus ≤1.5–8 μU/ml normal range), low or borderline 

low serum T4 levels (4.4 versus 4.5–11.5 μg/dL) and free-thyroxine indices (3.7 versus 3.8–10.8), and 

markedly elevated thyroid antimicrosomal antibody titers (1:6,400 or above). Serum T4 levels measured 

7 months earlier in two of the four men were normal. Antithyroglobulin antibodies were elevated in one 

of the four subjects (not evaluated in other workers). The exposed cohort had significantly more subjects 

with elevated serum thyrotropin (p=0.006), but free thyroxine index (p=0.06), serum T4 level (p=0.11) 

and antimicrosomal antibody titer (p=0.06) did not differ significantly from the controls. Questioning 
PBBs and PBDEs 75 

5. HEALTH EFFECTS 

about exposure to 74 occupational toxicants showed that three of the four hypothyroid subjects had only 

three common chemical exposures (PBBs, decaBDE, and bromine); the fourth worker was hired after 

PBB production ceased and was exposed only to decaBDE and bromine, but it is not clear if PBBs were 

still present in the work environment. Except for one control subject who had an enlarged thyroid, none 

of the exposed or control subjects had signs of thyroid enlargement, thyroid nodularity or hypothyroidism 

on physical examination, or had reported taking thyroid medication or having thyroid problems within the 

previous 5 years. Reevaluation of three of the four subjects 1 year later (none had been treated with 

thyroid hormone) showed that two still had low free-thyroxine indices and high serum thyrotropin, one 

had a normal free-thyroxine index and a high-normal serum thyrotropin, and all three still had markedly 

elevated thyroid antimicrosomal antibody titers. The findings of this study suggest that occupational 

exposure to PBBs, decaBDE, and/or bromine affected the thyroid, but the mixed chemical exposure and a 

lack of data on serum or tissue levels of the chemicals preclude attributing effects solely to any particular 

congener or mixture of congeners. 

Polybrominated Diphenyl Ethers. There is suggestive evidence of hypothyroidism in a small group of 

workers who were occupationally exposed to decaBDE as well as PBBs (Bahn et al. 1980), as 

summarized in the preceding subsection on endocrine effects of PBBs. In another study, plasma levels of 

thyroid hormones (T3 and free T4) and eight PBDE congeners (tetra- to heptaBDEs) were monitored for 

198–221 days in three electronic dismantling workers (Pettersson et al. 2002). The hormones remained 

within normal ranges and there were no correlations between levels of hormones and congeners. 

Inhalation studies of commercial octaBDE dust in rats showed no histopathological changes in the 

thyroids, parathyroids, adrenals, or pituitary following chamber exposure to 174 mg/m3 as powdered dust 

for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978), or in the adrenals 

(only endocrine tissue examined) following nose-only exposure to ≤250 mg/m3

 as dust aerosol for 

6 hours/day, 5 days/week for 14 days (Great Lakes Chemical Corporation 2000). Rats that were noseonly exposed to commercial octaBDE at levels of 1.1, 16, or 202 mg/m3

 for 6 hours/day, 5 days/week for 

13 weeks similarly showed no histological changes in the adrenals, pancreas, parathyroids, pituitary, or 

thyroids (Great Lakes Chemical Corporation 2001a, 2001b). Measurements of serum levels of thyroid 

hormones in the 13-week rat study, however, showed exposure-related decreases in mean thyroxine (total 

T4) at ≥16 mg/m3

 in both sexes, and increases in thyroid stimulating hormone (TSH) at ≥16 mg/m3 in 

males and 202 mg/m3

 in females. The changes were usually statistically significant (p<0.05 or p<0.01) 

compared to controls and were considered by the investigators to be consistent with chemical-induced 

hypothyroidism. There were no serum T3 changes, thyroid-attributable clinical signs or body weight 
PBBs and PBDEs 76 

5. HEALTH EFFECTS 

effects, or gross or histopathological changes in the thyroid. The 1.1 mg/m3 LOAEL for thyroid effects 

was used as the basis for the intermediate-duration MRL for inhalation exposure to octaBDE, as indicated 

in the footnote to Table 5-1 and discussed in Chapter 4 and Appendix A. 

Dermal Effects. 

Polybrominated Biphenyls. In a medical history survey study, 7 of 10 (70%) workers in the production 

department of a PBB manufacturing plant reported that they experienced symptoms of skin disorders, 

compared with 31% of 45 workers in other departments in the same plant and 18% in a control group of 

153 Wisconsin farm residents (Chanda et al. 1982). The survey covered a period of 3 years of potential 

exposure, but exposure levels were not reported. The dermatological symptoms were described as 

"almost uniformly" halogen acne (bromacne). Mean serum PBB levels for the respective PBB groups 

(with ranges listed in parentheses) were 603.9 ppb (11.4–1,729 ppb) and 16.5 ppb (4–234 ppb); PBBs 

were not detected in serum of the control subjects (Chanda et al. 1982). Physical examination confirmed 

the occurrence of bromacne in 13% of PBB workers compared with no acne in the control group. No 

other studies were located regarding dermal effects in humans after occupational exposure to PBBs. 

No studies were located regarding dermal effects in animals after inhalation exposure to PBBs. 

Polybrominated Diphenyl Ethers. No studies were located regarding dermal effects in humans after 

inhalation exposure to PBDEs. 

No gross or histological changes in the skin were observed in rats that were nose-only exposed to 

commercial octaBDE as dust aerosol at levels of ≤202 mg/m3

 for 6 hours/day, 5 days/week for 13 weeks 

(Great Lakes Chemical Corporation 2001a, 2001b). 

Ocular Effects. 

Polybrominated Biphenyls. No studies were located regarding ocular effects in humans after inhalation 

exposure to PBBs. 

Signs of ocular irritation (no further description) were observed in five male and five female rats that 

were exposed to a decabromobiphenyl dust mixture at 5,000 mg/m3

 concurrently with starch dust 

(6 mg/m3

) for 6 hours/day, 5 days/week for 4 weeks (Millischer et al. 1980). The seriousness of this 
PBBs and PBDEs 77 

5. HEALTH EFFECTS 

effect is unclear as severity was not reported and recovery was not assessed. Ocular irritation was not 

observed at 500 mg/m3

 and lower concentrations. 

Polybrominated Diphenyl Ethers. No studies were located regarding ocular effects in humans after 

inhalation exposure to PBDEs. 

Transient signs of ocular irritation that included eye squint, erythema, and/or ocular discharge were 

observed in rats that were chamber-exposed to pentaBDE aerosol (compound dissolved in corn oil), 

octaBDE dust, or decaBDE dust in concentrations of 2,000, 2,000, and 48,200 mg/m3

, respectively, for 

1 hour (IRDC 1974, 1975a, 1975b). Confidence in these effect levels is low due to a small number of 

tested animals and lack of control data. 

No histopathological changes were observed in eyes of rats that were chamber-exposed to ≤174 mg/m3

 of 

commercial octaBDE as powdered dust for 8 hours/day for 14 consecutive days (Great Lakes Chemical 

Corporation 1978). Opthalmoscopic and histological examinations showed no ocular effects in rats 

following nose-only exposure to ≤202 mg/m3

 of commercial octaBDE dust aerosol for 6 hours/day, 

5 days/week for 13 weeks (Great Lakes Chemical Corporation 2001a, 2001b). 

5.2.1.3 Immunological and Lymphoreticular Effects 

Polybrominated Biphenyls. Several immunological parameters were evaluated in a group of 28 workers 

from the Michigan Chemical Company who were involved in manufacturing and distributing PBBs 

(Stross et al. 1981). This company manufactured the FireMaster FF-1 that was involved in the 

agricultural contamination episode in Michigan in 1973–1974. The subjects had worked directly with 

PBBs during the previous 5 years, but exposure levels were not reported. Immunological analyses 

included determination of immunoglobulin levels, skin testing, and lymphocyte transformation studies. 

No abnormalities in lymphocyte number or function could be determined when compared to an 

unexposed group. One of three blastogenic responses (pokeweed mitogen [PWM]) was significantly 

reduced (p<0.01) relative to concurrent controls, but was within the normal control range for the 

laboratory. PWM is a mitogenic lectin that stimulates both human T and B cells. No specific information 

was provided regarding the skin testing and immunoglobulin levels. 

No studies were located regarding immunological or lymphoreticular effects in animals after inhalation 

exposure to PBBs. 
PBBs and PBDEs 78 

5. HEALTH EFFECTS 

Polybrominated Diphenyl Ethers. No studies were located regarding immunological or lymphoreticular 

effects in humans after inhalation exposure to PBDEs. 

No histopathological changes were observed in lymph nodes or bone marrow from rats that were exposed 

to 174 mg/m3

 of octaBDE dust for 8 hours/day for 14 consecutive days (Great Lakes Chemical 

Corporation 1978). 

Inhalation studies of commercial octaBDE dusts in rats showed no histopathological changes in the 

spleen, mesenteric or mediastinal lymph nodes, bone marrow, or spleen following chamber exposure to 

174 mg/m3

 as powdered dust for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 

1978), or in the spleens (only lymphoreticular tissue reported) following nose-only exposure to 

≤250 mg/m3

 as dust aerosol for 6 hours/day, 5 days/week for 14 days (Great Lakes Chemical Corporation 

2000). Rats that were nose-only exposed to commercial octaBDE at levels of 1.1, 16, or 202 mg/m3

 as 

dust aerosol for 6 hours/day, 5 days/week for 13 weeks similarly showed no effects in bone marrow, 

spleen, or thymus, although gross changes in pulmonary lymph nodes were observed at 202 mg/m3

. The 

effects included discolored and/or enlarged bronchial and mediastinal lymph nodes, and appeared to be 

associated with concurrent granulomatous inflammation of the lungs. 

5.2.1.4 Neurological Effects 

Polybrominated Biphenyls. Twenty-five workers at a PBB-manufacturing plant (exposure duration and 

levels not reported) displayed mean scores on tests of memory and learning that were typical for people of 

their age, and educational, occupational, and cultural backgrounds, even though they had an elevated 

mean PBB concentration in adipose tissue (9.33 ppm) (Brown et al. 1981). Workers with the highest 

concentrations of PBBs in adipose tissue showed no evidence of memory dysfunction in these tests. 

No studies were located regarding neurological effects in animals after inhalation exposure to PBBs. 

Polybrominated Diphenyl Ethers. No studies were located regarding neurological effects in humans 

after inhalation exposure to PBDEs. 

No clinical signs of neurotoxicity were observed in rats that were exposed to dusts of commercial 

octaBDE products at levels of 174 mg/m3

 for 8 hours/day for 14 consecutive days (Great Lakes Chemical 
PBBs and PBDEs 79 

5. HEALTH EFFECTS 

Corporation 1978), or ≤202 mg/m3 for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical 

Corporation 2001a, 2001b). Histological examinations of nervous system tissues, performed only in the 

13-week study, showed no effects in the brain (forebrain, midbrain, hindbrain), optic nerve, or a 

peripheral nerve (sciatic). 

5.2.1.5 Reproductive Effects 

Polybrominated Biphenyls. Eleven workers in a PBB manufacturing company (exposure duration and 

levels not reported) displayed no differences in the distribution of sperm counts, motility, or sperm 

morphology compared with a control group of 52 nonexposed men (Rosenman et al. 1979). PBBs were 

detected in the serum of all exposed subjects and in only one nonexposed subject, but no mean or 

individual serum PBB values were reported. 

No studies were located regarding reproductive effects in animals after inhalation exposure to PBBs. 

Polybrominated Diphenyl Ethers. No studies were located regarding reproductive effects in humans 

after inhalation exposure to PBDEs. 

No histopathological changes were observed in testes or ovaries from rats that were exposed to 

commercial octaBDE at concentrations of ≤174 mg/m3

 as powdered dust for 8 hours/day for 

14 consecutive days (Great Lakes Chemical Corporation 1978), or ≤250 mg/m3

 as dust aerosol for 

6 hours/day, 5 days/week for 2 weeks (Great Lakes Chemical Corporation 2001a, 2001b). A histological 

effect in the ovaries was found in a study of rats that were nose-only exposed to 0, 1.1, 16, or 202 mg/m3

as dust aerosol for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2001a, 

2001b). Absence of corpora lutea, based on qualitative evaluation of step sections of the ovary, was 

found in 3/10 females at 202 mg/m3

, compared to 0/10 incidences in the control and both lower exposure 

groups. The investigators interpreted this 30% increase in incidence be treatment-related because an 

absence of corpora lutea was considered unusual in rats at 20 weeks of age. No gross or histopathological 

changes were observed in the oviduct, uterus, or vagina, or in male reproductive tissues (testes with 

epididymides and vas deferens). 
PBBs and PBDEs 80 

5. HEALTH EFFECTS 

5.2.1.6 Developmental Effects 

No studies were located regarding developmental effects in humans or animals after inhalation exposure 

to PBBs or PBDEs. 

5.2.1.7 Cancer 

No studies were located regarding cancer in humans or animals after inhalation exposure to PBBs or 

PBDEs. 

5.2.2 Oral Exposure 

The highest NOAEL and all LOAEL values from each reliable study of health effects end points in each 

species and duration category for PBDEs are recorded in Tables 5-2 (PBBs), 5-3 (lower PDBEs), or 5-4 

(decaBDE) and plotted in Figures 5-2 (PBBs), 5-3 (lower PDBEs), or 5-4 (decaBDE). 

5.2.2.1 Death 

Polybrominated Biphenyls. No studies were located regarding death in humans after oral exposure to 

PBBs. 

Limited information is available on lethal amounts of PBBs in animals. In general, dosing regimen and 

magnitude affect response. The lack of decreased survival in some studies does not necessarily indicate 

low toxicity because observation periods may not be sufficient to observe effects that develop slowly. 

Except as noted below, acute-duration studies administered PBBs by gavage in oil vehicle. A single 

1,000 mg/kg dose of FireMaster FF-1 did not significantly increase mortality in rats observed for ≤2 years 

posttreatment (Kimbrough et al. 1978b, 1981). Exposing pregnant rats to ≤800 mg/kg FireMaster BP-6 

on one of gestation days 6–14 did not significantly increase mortality, but the animals were not observed 

beyond pregnancy (Beaudoin 1977). Administration of 1,000 mg/kg/day FireMaster FF-1 for 6–10 doses 

(5 days/week), however, caused 100% mortality in rats; the mean time to death was 12.3 days in females 

and 11.0 days in males (Gupta and Moore 1979). The cause of death was not specifically reported, but a 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

ACUTE EXPOSUREDeath

1

286

1000

1000 (18/18 died)

2 wk5d/wk1x/d(GO)

(Fischer344/N)Rat 

2

19

130

130 F (63% lethality)

14 dad lib(F)

(Balb/c)Mouse 

Systemic

3

286

Endocr 286 M

55

once

(Sprague- (GO)Dawley)Rat 

4

1

1

b

Endocr

58

3

3 M (decreased thyroid plasma T4

hormone)

10 d1x/d(GO)

(SpragueDawley)Rat 

5

400

Bd Wt 400 F

50

800

800 F (unknown percent maternal

weight loss)

onceGd 6-14(GO)

(Wistar)Rat 

6

Hepatic

99

500

500 M (increased hepatic

phospholipids and serumcholesterol)

once18 mo observ(GO)

(Sherman)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

7

Hepatic

289

1000

1000 (hepatocytic swelling, fatty

infiltration, multinucleation,necrosis, and cytolysis)

2 wk5 d/wk1x/d(GO)

(Fischer344/N)Rat 

Renal

1000

1000 (darkened kidneys)

Endocr

1000

1000 (darkened adrenal glands)

Bd Wt

1000

1000 (unknown percent weight loss,

emaciation)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

8

30

Resp 30

310

2 wk5 d/wk1x/d(GO)

(Fischer344/N)Rat 

30

Cardio 30

30

Gastro 30

30

Hemato 30

30

Musc/skel 30

0.3

Hepatic 0.3

3

3 (dose-related hepatocyte

enlargement and single-cellnecrosis)

30

Renal 30

30

Endocr 30

30

Dermal 30

30

Ocular 30

30

Bd Wt 30
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

9

1000

Resp 1000

206

once2-14 mo observ(GO)

(Sherman)Rat 

1000

Cardio 1000

1000

Gastro 1000Hepatic

1000

1000 (vacuolation, necrosis, and

fibrosis, porphyria,multinucleation)

1000

Renal 1000

1000

Endocr 1000

1000

Bd Wt 1000

10

Hepatic

184

500

500 M (vacuolation of hepatocytes)

once15wk observ(GO)

(Sherman)Rat 

500

Bd Wt 500 M
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

11

200

Resp 200 F

187

once18-22mo observ(GO)

(Sherman)Rat 

200

Cardio 200 FHepatic

200

200 F (porphyrin accumulation)

200

Renal 200 F

200

Endocr 200 F

200

Dermal 200 F

200

Bd Wt 200 F
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

12

1000

Resp 1000 F

198

once23mo observ(GO)

(Sherman)Rat 

1000

Cardio 1000 F

1000

Gastro 1000 F

1000

Musc/skel 1000 FHepatic

1000

1000 F (hepatomegaly, hepatocyte

enlargement and vacuolation,porphyrin accumulation)

1000

Renal 1000 F

1000

Endocr 1000 F

1000

Ocular 1000 FBd Wt

1000

1000 F (12% decreased body weight

gain)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

13

Hepatic

93

1000

1000 M (fatty changes in centrilobular

hepatocytes)

once28d observ(GO)

(SpragueDawley)Rat 

1000

Bd Wt 1000 M

14

Hepatic

94

3000

3000 M (fatty changes in centrilobular

hepatocytes)

2 d1x/d(GO)

(SpragueDawley)Rat 

1000

Bd Wt 1000 M

15

2000

Bd Wt 2000 F

165

once

(Sprague- (GO)Dawley,Spartan)Rat 

16

Hepatic

179

5

5 M (hepatomegaly and fatty

changes in weanlings)

10 dad lib(F)

(Fischer 344)Rat 

5

Bd Wt 5 M
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

17

0.66

Hepatic 0.66 M

102

6.53

6.53 M (hyperplasia and fatty changes)

2 wkad lib(F)

(SpragueDawley)Rat 

71

Renal 71 M

71

Endocr 71 M

71

Bd Wt 71 M

18

36

Hepatic 36 F

139

2 wkad lib(F)

(SwissWebster)Mouse 

19

Hepatic

47

130

130 F (focal areas of coagulative

necrosis)

11 dad lib(F)

(Swiss/ IRC)Mouse 

20

Bd Wt

96

130

130 M (30% decreased body weight)

4-14 dad lib(F)

(Swiss/ IRC)Mouse 

21

Bd Wt

21

130

130 F (23% weight loss)

14 dad lib(F)

(Balb/c)Mouse 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

22

30

Resp 30

306

2 wk5 d/wk1x/d(GO)

(B6C3F1)Mouse 

30

Cardio 30

30

Gastro 30

30

Hemato 30

30

Musc/skel 30

0.3

Hepatic 0.3

3

3 (dose-related increase in

incidence of hepatocyteenlargement and single-cellnecrosis)

30

Renal 30

30

Endocr 30

30

Ocular 30

30

Bd Wt 30

30

Other 30
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Immuno/ Lymphoret

23

287

1000

1000 (atrophy of thymus; necrosis of

splenic lymphoblasts)

10 d5 d/wk1x/d(GO)

(Fisher344/N)Rat 

24

20

130

130 F (suppressed antibody-mediated

response to SRBC, thymicatrophy)

14 dad lib(F)

(Balb/c)Mouse 

Reproductive

25

196

1000

1000 F (9% increased incidence of

uterine polyps)

once23mo observ(GO)

(Sherman)Rat 

26

21

21 F

73

63

63 F (29% reduction in success of

pregnancy)

9 dGd 6-151x/d(F)

(C57BL)Mouse 

Developmental

27

40

40

49

200

200 (9.1-31.4% resorptions)

onceGd 6-14(GO)

(Wistar)Rat 

28

5

5

43

50

50 (12% decrease in fetal body

weight)

14 dGd 7-20(F)

(SpragueDawley)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

29

160

200

200 (increased mortality and liver

neoplasms in offspring)

Gd 7-14 1x/d

(Sherman) (GO)Rat 

30

42

42.9

42.9 (12-20% decreased mean body

weight in treated pups atpost-parturition day 60)

9 dGd 7-151x/d(GO)

(NS)Rat 

31

1000

1000

399

10 dGd 6-151x/d(GO)

(SpragueDawley, Iffacredo)Rat 

32

9.1

9.1

264

86

86 (increased incidence of extra

ribs)

10 dGd 6-15ad lib(F)

(ChR-CD)Rat 

33

5

5

45

50

50 (cleft palate)

12 dGd 7-18(F)

(Swiss/ IRC)Mouse 

Cancer

34

161

200

200 (CEL: hepatocellular carcinoma

in offspring)

Gd 7-14 1x/d

(Sherman) (GO)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

35

197

1000

1000 F (CEL: hepatocellular carcinoma

once

(Sherman) (GO)Rat 

INTERMEDIATE EXPOSUREDeath

36

290

149

149 M (90-day LD50)

65

65 F (90-day LD50)

c

4.5 wk5 d/wk22 d(GO)

(Fischer344/N)Rat 

37

144

0.3

0.3 M (decreased mean survival time)

25 wk5 d/wk1x/d(GO)

(Fischer344/N)Rat 

38

150

10

10 F (decreased mean survival time)

25 wk5 d/wk1x/d(GO)

(B6C3F1)Mouse 

39

261

4

4 M (4/6 died)

30 dad lib(F)

(NS)Gn Pig 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

40

128

2

2 F (7/8 died)

45 dad lib(F)

(NS)Gn Pig 

41

250

0.47

0.47 M (LD50)

c

0.61

0.61 F (LD50)

313 dad lib(F)

(NS)Mink 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Systemic

42

Gastro

285

18

18 F (hyperplastic gastroenteritis)

137 dad lib(F)

(Rhesus)Monkey 

Hemato

18

18 F (decreased RBC, PCV, and

WBC)

Hepatic

18

18 F (enlarged hepatocytes,

hyperplasia of bile ductepithelium, increased SGPT,decreased serum cholesterol)

Renal

18

18 F (hyperplasia of bladder

epithelium)

Endocr

18

18 F (adrenal hemorrhage)

Dermal

18

18 F (edema, atrophy, and

squamous metaplasia ofsebaceous glands)

Bd Wt

18

18 F (27% body weight loss)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

43 Cardio

276

0.73

0.73 M (enlarged heart at necropsy)

25-50 wkad lib(F)

(Rhesus)Monkey 

Gastro

0.73

0.73 M (proliferation of mucosal cells,

chronic inflammatory cells,severe ulcerative colitis)

Hemato

0.73

0.73 M (decreased PCV and total

serum protein)

Hepatic

0.73

0.73 (enlarged hepatocytes with

increased lipid droplets,increased SGPT, decreasedserum cholesterol, hyperplasiaof bile duct epithelium)

Dermal

0.73

0.73 (edema and alopecia,

keratinization of hair folliclesand sebaceous glands)

Bd Wt

0.73

0.73 (34% weight loss in adult male,

0% weight gain in juvenile)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

44

10

Resp 10 M

54

30 dad lib(F)

(SpragueDawley)Rat 

10

Cardio 10 M

10

Gastro 10 M

10

Hemato 10 M

10

Musc/skel 10 MHepatic

0.1

0.1 M (hepatocyte swelling,

vacuolation)

10

Renal 10 M

10

Endocr 10 M

10

Dermal 10 M

10

Ocular 10 M

10

Bd Wt 10 M
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

45

Endocr

98

0.05

0.05 M (altered thyroid follicular

ultrastructure)

30 dad lib(F)

(SpragueDawley)Rat 

46

Endocr

56

1

1 M (decreased serum thyroid

hormone T4)

20 d1x/d(GO)

(SpragueDawley)Rat 

47

Endocr

32

0.45

0.45 F (decreased thyroid serum T3

and T4 hormones)

7 moad lib(F)

(SpragueDawley)Rat 

48

0.05

Endocr 0.05 F

35

0.25

0.25 F (decreased adrenal serum

corticosterone B, DHEA andDHS hormones)

5-7 moad lib(F)

(SpragueDawley)Rat 

49

6

Endocr 6 M

34

20 d1x/d(GO)

(SpragueDawley)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

50

2

Resp 2 M

342

28 dad lib(F)

(SpragueDawley)Rat 

2

Cardio 2 M

2

Gastro 2 M

2

Hemato 2 MHepatic

2

2 M (increased liver weight,

increased liver microsomalenzymes, fatty degeneration ofliver)

2

Renal 2 MEndocr

2

2 M (reduction of follicular size and

colloid density and exfoliation ofepithelium in thyroid)

2

Dermal 2 M

2

Bd Wt 2 M
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

51

Hepatic

221

0.5

0.5 M (bile duct hyperplasia)

82 dad lib(F)

(SpragueDawley)Rat 

Ocular

5

5 M (xerophthalmia)

5

Bd Wt 5 M
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

52

1000

Cardio 1000

294

4.5 wk5 d/wk1x/d(GO)

(Fischer344/N)Rat 

1000

Gastro 1000Hemato

30

30 (decreased hemoglobin, PCV,

and platelet count)

Hepatic

30

30 (hepatocyte enlargement, fatty

infiltration and multinucleation,porphyrin accumulation)

Renal

30

30 (dilation of Bowman's capsule

with serous fluid)

Endocr

30

30 (unspecified altered thyroid

histology)

Bd Wt

30

30 (19% decreased body weight

gain) 100

100 (emaciation)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

53

30

Resp 30

298

30 d5 d/wk1x/d(GO)

(Fischer344/N)Rat 

30

Cardio 30

30

Gastro 30

30

Hemato 30

30

Musc/skel 30

0.3

Hepatic 0.3

3

3 (increased liver weight,

hepatocyte swelling, andnecrosis)

30

Renal 30

3

Bd Wt 3

30

30 (significant decrease in body

weight)

54

Endocr

31

0.25

0.25 M (colloid droplets, abnormal

microvilli and other changes inthyroid follicle ultrastructure)

5 wkad lib(F)

(Holtz- man)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

55

0.25

Hepatic 0.25 M

97

2.5

2.5 M (hepatocyte hypertrophy and

degeneration)

5 wkad lib(F)

(Holtz- man)Rat 

56

Hemato

12

10

10 F (increased white blood cell

count)

6 mo5 d/wk(GO)

(Fischer 344)Rat 

1

Bd Wt 1 F

3

3 F (15% decreased weight gain)

57

Hepatic

178

5

5 F (enlarged and vacuolated

hepatocytes, focal necrosis)

3 moad lib(F)

(SpragueDawley)Rat 

Renal

5

5 F (degenerative changes in

glomeruli)

Bd Wt

5

5 F (10% decreased body weight

gain)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

58

100

Gastro 100

397

13 wk

(Sprague- (F)Dawley CFY)Rat 

100

Hemato 100

25

Hepatic 25

100

100 (11% increased liver weight,

hepatocyte hypertrophy andvacuolization, slightly increasedliver lipids)

100

Renal 100

100

Bd Wt 100
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

59

800

Cardio 800 M

173

30 dad lib(F)

(SpragueDawley)Rat 

80

Hemato 80 M

800

800 M (decreased PCV and RBC

counts)

Hepatic

8

8 M (enlargement and vacuolation)

Renal

8

8 M (hyaline degenerative changes)

Endocr

8

8 M (thyroid hyperplasia)

800

Bd Wt 800 M
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

60

1

Cardio 1

177

8 moad lib(F)

(SpragueDawley)Rat 

1

Hemato 1

1

Hepatic 1

1

Renal 1

1

Endocr 1

1

Bd Wt 1
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

61

10

Resp 10

149

25 wk5 d/wk1x/d(GO)

(Fischer344/N)Rat 

10

Cardio 10

0.3

Gastro 0.3

1

1 (gastric ulcers)

0.1

Hemato 0.1

0.3

0.3 (decreased hemoglobin, MCH,

PCV, and MCV)

10

Musc/skel 10

0.1

Hepatic 0.1

0.3

0.3 (lipid accumulation; increased

atypical foci; porphyrinaccumulation)

0.3

Renal 0.3

1

1 (chronic progressive

nephropathy)

0.1

Endocr 0.1

0.3

0.3 (decreased serum thyroid T4

hormone)

10

Ocular 10

10

Bd Wt 10
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

62

2.5

Hepatic 2.5 F

59

7 moad lib(F)

(SpragueDawley)Rat 

2.5

Endocr 2.5 F

63

2.5

Endocr 2.5 F

60

7 moad lib(F)

(SpragueDawley)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

64

50

Resp 50 M

259

30 dad lib(F)

(SpragueDawley)Rat 

50

Cardio 50 M

50

Gastro 50 M

50

Hemato 50 M

50

Musc/skel 50 MHepatic

1

1 M (hepatocyte vacuolation)

10

Renal 10 M

50

50 M (unquantified, but significantly

increased BUN)

50

Endocr 50 M

10

Bd Wt 10 M

50

50 M (16% decreased weight gain)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

65

5

Resp 5 M

158

30 dad lib(F)

(SpragueDawley)Rat 

5

Cardio 5 M

5

Gastro 5 M

5

Hemato 5 M

0.5

Hepatic 0.5 M

5

5 M (hepatocyte swelling and

vacuolation)

5

Renal 5 M

0.5

Endocr 0.5 M

5

5 M (hyperplasia of thyroid follicular

epithelium)

0.5

Bd Wt 0.5 M

5

5 M (27-36% reduced body weight

gain)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

66

0.66

Hepatic 0.66 M

107

6.53

6.53 M (hyperplasia and progressive

lipid changes)

4 wkad lib(F)

(SpragueDawley)Rat 

71

Renal 71 M

71

Endocr 71 M

71

Bd Wt 71 M

67

13

Bd Wt 13 F

23

30 dad lib(F)

(Balb/c)Mouse 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

68

30

Resp 30

302

30 d5 d/wk1x/d(GO)

(B6C3F1)Mouse 

30

Cardio 30

30

Gastro 30Hemato

30

30 F (decreased PCV)

30

Musc/skel 30

0.3

Hepatic 0.3

3

3 (hepatocyte enlargement and

necrosis)

30

Renal 30

30

Ocular 30

3

Bd Wt 3

30

30 M (significant decrease in weight

gain)

30

Other 30
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

69

21.7

Resp 21.7 M

66

6 wkad lib(F)

(Balb/ cBYJ)Mouse 

21.7

Cardio 21.7 M

0.65

Hepatic 0.65 M

21.7

21.7 M (hepatocellular necrosis and

vacuolation)

21.7

Renal 21.7 M

0.65

Endocr 0.65 M

0.65

Bd Wt 0.65 M

21.7

21.7 M (33% reduction in body weight)

70

10

Hemato 10

14

6 mo5 d/wk(GO)

(B6C3F1)Mouse 

Endocr

10

10 (increased adrenal weight gain)

10

Bd Wt 10
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

71

10

Resp 10

155

25 wk5 d/wk1x/d(GO)

(B6C3F1)Mouse 

10

Cardio 10

10

Gastro 10

0.1

Hemato 0.1

0.3

0.3 (decreased erythrocyte count

and MCV)

10

Musc/skel 10

0.1

Hepatic 0.1

0.3

0.3 (increased liver weight, SGOT

and porphyrin accumulation)

10

Renal 10

10

Endocr 10

10

Ocular 10

3

Bd Wt 3 M

10

10 M (25% decreased body weight

gain)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

72

20

Resp 20 M

262

30 dad lib(F)

(NS)Gn Pig 

Hepatic

0.04

0.04 M (vacuolation and fatty changes)

0.4

Bd Wt 0.4 M

4

4 M (severe weight loss prior to

death)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

73

8

Cardio 8

182

16 wkad lib(F)

(NS)Pig 

1

Gastro 1

8

8 (gross hyperplasia glandular

stomach)

8

Hemato 8Hepatic

1

1 (LDH increased)

8

Renal 8Endocr

8

8 (increased adrenal weight)

1

Dermal 1

8

8 (dermatosis)

Bd Wt

1

1 (12.9% reduced weight gain and

food intake)

74

0.125

Hepatic 0.125 F

10

1.25

1.25 F (fatty changes and necrosis)

12 wkGwk 8-ppwk 4ad lib(F)

(NS)Pig 

1.25

Endocr 1.25 F

2.5

2.5 F (significant decrease in thyroid

serum T3 and T4 hormones)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

75

2.4

Cardio 2.4

255

313 dad lib(F)

(NS)Mink 

Hepatic

0.24

0.24 F (48% increased relative liver

weight, fatty infiltration)

2.4

Renal 2.4Bd Wt

0.39

0.39 F (14% decreased prebreeding

body weight gain)

1.86

1.86 F (up to 19% mean body weight

loss prior to death)

Immuno/ Lymphoret

76

0.03

0.03 M

24

3

3 M (decreased lymphocytic

response to mitogenstimulation; decrease inabsolute and relative thymusweight)

5 wk5 d/wk(GO)

(Fischer 344)Rat 

77

1

1 F

11

3

3 F (decreased lymphoproliferative

responses and decreaseddelayed hypersensitivityresponses)

6 mo5 d/wk1x/d(GO)

(Fischer 344)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

78

0.13

0.13 F

22

1.3

1.3 F (reduced antibody mediated

response to SRBC and 21%reduction in thymus weight)

30 dad lib(F)

(Balb/c)Mouse 

79

0.65

0.65 M

65

21.7

21.7 M (increased lethality due to

endotoxin challenge)

6 wkad lib(F)

(Balb/ cBYJ)Mouse 

80

3

3

13

10

10 (increased lethality due to

infection with L monocytogenesdecreased response to mitogenstimulation)

6 mo5 d/wk1x/d(GO)

(B6C3F1)Mouse 

81

129

0.4

0.4 F (reduced antitoxin titers

following toxoid challenge)

4

4 F (thymic atrophy and follicular

depletion in spleen)

45 dad lib(F)

(NS)Gn Pig 

82

1.25

1.25 F

110

2.5

2.5 F (reduced lymphocyte response

to mitogen stimulation)

12 wkGwk 8- ppwk 4ad lib(F)

(NS)Pig 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Neurological

83

74

10

10 (decreased limb strength)

6 mo5 d/wk1x/d(GO)

(NS)Rat 

84

3

3 M

162

6

6 M (decreased motor activity)

4 wk5 d/wk1x/d(G)

(SpragueDawleyHoltz- man)Rat 

85

0.2

0.2

1002

2

2 (delayed acquisition of

locomotion and reduced openfield activity in offspring)

40dGd 6-Ppd 24(F)

(SpragueDawley)Rat 

86

88

3

3 (decreased motor activity, grip

strength, and startleresponsiveness)

6 mo5 d/wk(GO)

(Fischer344/N)Rat 

87

89

30

30 (decreased open field motor

activity and grip strength)

4 wk5 d/wk1x/d(GO)

(Fischer344/N)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Reproductive

88

275

0.73

0.73 M (hypoactive seminiferous

tubules)

25-50 wkad lib(F)

(Rhesus)Monkey 

89

246

0.012

0.012 F (increased menstrual cycle

duration in 4/7; implantationbleeding in 2/7)

6 moad lib(F)

(Rhesus)Monkey 

90

14.3

14.3 F

137

28.6

28.6 F (no implantations in 2/5 rats)

15 dGd 0-148x(GO)

(Wistar)Rat 

91

30

30 M

c

1000

1000 F

293

100

100 M (squamous metaplasia,

hyperplasia, and necrosis inepithelium of ductus deferens)

4-5 wk5 d/wk22 d(GO)

(Fischer344/N)Rat 

92

5

5 F

142

42 dGd 8-ppd 28ad lib(F)

(SpragueDawley)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

93

30

30

301

4-5 wk5 d/wk1x/d(GO)

(B6C3F1)Mouse 

94

0.24

0.24

254

0.39

0.39 F (10% reduction in body weight)

313 dad lib(F)

(NS)Mink 

Developmental

95

2.9

2.9

136

14.3

14.3 (increased resorptions)

15 dGd 0-148x(GO)

(Wistar)Rat 

96

5

5

109

4 wk5 d/wk(G)

(Holtz- manSpragueDawley)Rat 

97

108

0.2

0.2 (deficits in learning behavior in

offspring, 6 months afterprenatal and lactationalexposure)

40 dGd 6- ppd24ad lib(F)

(SpragueDawley)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

98

1001

0.2

0.2 M (reduced crown-rump length)

40dGd 6-Ppd 24(F)

(SpragueDawley)Rat 

99

141

0.5

0.5 (increased liver weight,

hepatocyte vacuolation,decreased hepatic vitamin Acontent in F1 but not F2)

5

5 (decreased pup survival during

lactation in F1)

42 dGd 8-ppd 28ad lib(F)

(SpragueDawley)Rat 

100

67

5

5 (20% decrease in pup body

weight gain, 50% decreasedhepatic vitamin A, 256-285%decreased urinary uro- andcoproporphyrins in pups)

42-126 dGd 8-ppd 28-112ad lib(F)

(SpragueDawley)Rat 

101

61

2.5

2.5 (decreased body weight,

increased relative liver weight,and decreased serum T4 inoffspring)

37 dGd 0-ppd 15ad lib(F)

(SpragueDawley)Rat 

102

358

0.5

0.5 (hepatic vacuolization and

altered foci in pups)

77 dGd 0-ppd 56ad lib(F)

(Fischer344/N)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

103

2

2

15

3

3 (decreased hematocrit in

offspring)

10

10 (early postnatal death; no

details provided)

42 dGd 0- weaning1x/2d(GO)

(B6C3F1)Mouse 

104

371

1.5

1.5 (hepatic cytomegaly and altered

foci in pups)

77 dGd 0- ppd 56ad lib(F)

(B6C3F1)Mouse 

105

17

3

3 (performance deficits in

offspring in a learned task)

Gd 0-ppd 211x/2d(GO)

(C57B1/6)Mouse 

106

0.125

0.125

9

1.25

1.25 (increased relative liver weight,

decreased serum thyroidhormone levels, and slightthyroid hyperplasia in offspring)

12 wkGwk 8- ppwk 4ad lib(F)

(NS)Pig 

107

253

0.155

0.155 (decreased birth and 4-week

weights in kits)

313 d

(NS) (F)Mink 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Cancer

108

191

100

100 (CEL: hepatocellular carcinoma

4 mo12 x(GO)

(Sherman)Rat 

109

148

3

3 (CEL: hepatocellular carcinoma

25 wk5 d/wk1x/d(GO)

(Fischer344/N)Rat 

110

154

10

10 (CEL: hepatocellular

carcinomas)

25 wk5 d/wk1x/d(GO)

(B6C3F1)Mouse 

111

370

1.5

1.5 (CEL: hepatocullular adenoma

and carcinoma in offspring)

77 dGd 0-ppd 56ad lib(F)

(B6C3F1)Mouse 

CHRONIC EXPOSUREDeath

112

347

0.5

0.5 M (18% decreased survival)

115 wksGd 0-ppd 56 (weaning)104 wks post-weaningad lib(F)

(Fischer344/N)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

113

360

1.3

1.3 F (44% decreased survival)

116 wksGd 0-ppd 56 (weaning)105wks post-weaningad lib(F)

(B6C3F1)Mouse 

Systemic

114

0.012

Hemato 0.012

247

66 wkad lib(F)

(Rhesus)Monkey 

Bd Wt

0.012

0.012 (7.4% weight loss)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

115

1.5

Resp 1.5

352

115 wksGd 0-ppd 56;104 wks post-weaning(F)

(Fischer344/N)Rat 

1.5

Cardio 1.5

0.5

Gastro 0.5

1.5

1.5 M (forestomach hyperplasia,

inflammation, ulceration)

0.5

Hemato 0.5

1.5

1.5 F (mild anemia)

1.5

Musc/skel 1.5Hepatic

0.5

0.5 (hypertrophy, vacuolation,

altered foci, increased serumcholesterol, decreased serumtriglycerides)

1.5

Renal 1.5

1.5

Endocr 1.5

1.5

Dermal 1.5Bd Wt

0.5

0.5 (11-18% decreased final body

weight)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

116

3.9

Resp 3.9

365

116 wksGd 0-ppd 56 (weaning)105 wks post-weaningad lib(F)

(B6C3F1)Mouse 

3.9

Cardio 3.9

3.9

Gastro 3.9

3.9

Musc/skel 3.9Hepatic

1.3

1.3 (hypertrophy, vacuolization,

single-cell necrosis, altered foci,bile duct hyperplasia)

1.3

Renal 1.3

3.9

3.9 (increased chronic nephropathy

Endocr

1.3

1.3 (thyroid follicular cell

hyperplasia)

3.9

Dermal 3.9

3.9

Bd Wt 3.9
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Immuno/ Lymphoret

117

0.5

0.5 M

348

1.5

1.5 M (splenic fibrosis)

115 wksGd 0-ppd 56 (weaning)104 wks post-weaningad lib(F)

(Fischer344/N)Rat 

118

1.3

1.3

361

3.9

3.9 (increased splenic

hematopoiesis)

116 wksGd 0-ppd 56 (weaning)105 wks post-weaningad lib(F)

(B6C3F1)Mouse 

Reproductive

119

1.5

1.5 M

0.5

0.5 F

c

350

1.5

1.5 F (cystic endometrial hyperplasia)

115 wksGd 0-ppd-56 (weaning)104 wks post-weaningad lib(F)

(Fischer344/N)Rat 

120

3.9

3.9

363

116 wksGd 0-ppd 56 (weaning)105 wks post-weaningad lib(F)

(B6C3F1)Mouse 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Developmental

121

245

0.012

0.012 (1/7 fetuses were aborted,1/7

fetuses stillborn, 12%decreased birth weight and 22%decreased postnatal weight gainin 4/7 survivors)

359-469 dad lib(F)

(Rhesus)Monkey 

Cancer

122

351

1.5

1.5 (CEL: leukemia)

0.5

0.5 (CEL: hepatocellular adenoma

and carcinoma)

115 wksGd 0- ppd 56104 wks post-weaning(F)

(Fischer344/N)Rat 

123a The number corresponds to entries in Figure 5-2.b Used to derive an acute-duration oral minimal risk level (MRL) of 0.01 mg/kg/day. The MRL was derived by dividing the NOAEL by an uncertainty factor of 1from animals to humans, 10 for human variability.c Differences in levels of health effects and cancer effects between male and females are not indicated in Figure 5-2. Where such differences exist, only the lemost sensitive gender are presented.Ad lib - ad libitum; Bd Wt = body weight; BP-6 = FireMaster BP-6; Cardio = cardiovascular; CEL = cancer effect level; d = day(s); DBB = deca=brominated biph

water; endocr = endocrine; (F)= feed; F = female; FF-1 = FireMaster FF-1; (G) = gavage; gastro = gastrointestinal; Gd = gestation day; Gn Pig - Guinea Pig; (Gwk = gestation week; HBB = hexa-brominated biphenyl; hemato = hematological; hr = hour(s); LD50 = lethal dose; 50% kill; LOAEL = lowest observed advermale; min = minute(s); mo = month(s); Musc/skel = musculoskeletal; NOAEL = no observed adverse effect level; NS = not specified; OBB octa-brominated bipobservation; ppd = post partum day; ppwk = post partum week; Resp = respiratory; T4 = thyroxine; wk = week(s); x = time(s)

364

3.9

3.9 (thyroid follicular cell adenoma)

1.3

1.3 (CEL: hepatocellular adenoma

and carcinoma)

116 wksGd 0- ppd 56105wks post-weaning(F)

(B6C3F1)Mouse 
0.01

0.1

1

10

100

1000

10000

Death

2m1r

Respiratory

22m 8r

9r

11r

12r

Cardiovascular

22m 8r

9r

11r

12r

Gastrointestinal

22m 8r

9r 12r

Hematological

22m 8r

Musculoskeletal

22m 8r

12r

Hepatic

18m

19m

22m

22m

6r

7r

8r

9r

10r11r

12r 

17r

mg/kg/day

Figure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral

Acute (≤14 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic
0.01

0.1

1

10

100

1000

10000

Endocrine

22m3r

4r4r

7r

8r

9r

11r

12r

17r

Dermal

8r11r

Ocular

22m 8r

12r

Body Weight

20m 21m

22m

5r5r7r

8r

9r10r

11r

12r 13r 14r

15r

16r17r

Other

22m

Immuno/Lymphor

24m23r

Reproductive

26m26m25r

32

mg/kg/day

Figure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continued

Acute (≤14 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic

*Doses represent the lowest dose tested per study that produced a tumorigenicresponse and do not imply the existence of a threshold for the cancer endpoint.
0.01

0.1

1

10

100

1000

10000

Death

39g40g41n38m36r37r

Respiratory

72g68m

69m

71m 44r

50r

53r

61r64r65r

Cardiovascular

75n43k68m69m71m 44r

50r

52r

53r

59r

60r

61r64r65r73p

Gastrointestinal

42k

43k

68m

71m 44r

50r52r

53r

58r61r61r64r

65r

73p

73p

Hematological

42k

43k

68m

70m

71m71m

44r

50r52r 53r

56r

58r 59r59r

60r61r61r

64r

65r73p

mg/kg/day

Figure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continue

Intermediate (15-364 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic
0.01

0.1

1

10

100

1000

10000

Hepatic

55r55r57r

58r

58r59r60r61r

61r62r64r65r65r66r66r

73p 74p

74p

Renal

75n

42k68m

69m

71m 44r

50r

52r 53r

57r

58r

59r

60r 61r

61r

64r

64r

65r

66r

73p

mg/kg/day

Figure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continue

Intermediate (15-364 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Endocrine

42k

69m

70m 71m 44r

45r46r47r

48r

48r

49r50r52r

54r

59r

60r

61r

61r62r 63

6465656673p74p

72g

75n

42k

43k68m68m

69m

69m71m71m

44r50r51r

52r

53r53r

Systemic
0.01

0.1

1

10

100

1000

10000

Body Weight

72g72g75n

75n

42k

43k

67m

68m

68m

69m

69m

70m 71m

71m

44r

50r51r

52r

52r

53r

53r

56r

56r57r

58r

59r

60r

61r 64r

64r

65r

65r

66r

73p

Other

68m

mg/kg/day

Figure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continued)

Intermediate (15-364 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Dermal

42k43k44r50r73p73p

Ocular

68m71m 44r

51r

61r

Systemic
0.01

0.1

1

10

100

1000

10000

Developmental

95r96r97r 98r

99r99r

100r101r102r106p106p

Cancer *

110m111m108r109r

mg/kg/day

Figure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continued)

Intermediate (15-364 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Neurological

83r84r84r85r85r

86r87r

Reproductive

94n94n88k89k93m 90r

90r

91r91r92r

107n103m103m103m104m

105m95r

Immuno/Lymphor

81g81g78m78m79m79m80m80m76r

76r77r

77r82p

82p

*Doses represent the lowest dose tested per study that proresponse and do not imply the existence of a threshold for 
0.01

0.1

1

10

Death

113m

112r

Respiratory

116m115r

Cardiovascular

116m115r

Gastrointestinal

116m115r115r

Hematological

114k

115r115r

Musculoskeletal

116m115r

Hepatic

116m115r

Renal

116m116m115r

Endocrine

116m115r

Dermal

116m115r

Body Weight

114k116m115r

Immuno/Lymphor

118m118m117r117r

Rep

121111

mg/kg/day

Figure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Contin

Chronic (≥365 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic

*Doses represent the lowest dose tested per study that produced a tumorigenicresponse and do not imply the existence of a threshold for the cancer endpoint.
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral

(mg/kg) (mg/kg) (mg/kg) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

ACUTE EXPOSUREDeath

1

2106

6200

6200 (44-day LD50)

once

(Wistar) (GO)Rat 

2

2090

5000

5000 (4/5 died)

once

Spartan (GO)Rat 

3

2107

5000

5000 (14-day LD50)

once

(Sprague- (GO)Dawley)Rat 

Systemic

4

10

Bd Wt 10

2071

100

100 (30% reduced maternal body

weight gain)

10 dGd 6-15(GO)

(CD)Rat 

5

76.6

Hepatic 76.6 M

2042

14 d1x/d(GO)

(SpragueDawley)Rat 

6

56.4

Hepatic 56.4 M

2043

14 d1x/d(GO)

(SpragueDawley)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

7

Endocr

2109

18

18 (reduced serum T4)

14 d1x/d(GO)

(SpragueDawley)Rat 

8

Hepatic

2068

18

18 (reduced liver vitamin A)

14 d1x/d(GO)

(SpragueDawley)Rat 

Endocr

18

18 (reduced serum T4)

9

5000

Bd Wt 5000

2089

once

Spartan (GO)Rat 

10

500

Bd Wt 500

2091

once

Spartan (GO)Rat 

11

25

Bd Wt 25

2075

10 dGd 6-15(GO)

(CD)Rat 

12

25

Bd Wt 25

2073

50

50 (40% reduced maternal body

weight gain)

10 dGd 6-15(GO)

(CD)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

13

10

Endocr 10

2081

30

30 (reduced serum T4)

4 d1x/d(GO)

(LongEvans)Rat 

14

3

Endocr 3

2082

10

10 (reduced serum T4)

4 d1x/d(GO)

(LongEvans)Rat 

15

Endocr

2110

18

18 (reduced serum T4)

14 d1x/d(GO)

(C57BL/6N)Mouse 

16

500

Hepatic 500

2059

once

C57BL/6J (GO)Mouse 

100

Endocr 100

500

500 (reduced serum T4)

17

72

Hepatic 72

2061

14d1x/d(GO)

(C57BL/6N)Mouse 

Endocr

18

18 (reduced serum T4)

72

Bd Wt 72
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

18

18

Hepatic 18

2069

36

36 (reduced liver vitamin A)

14 d1x/d(GO)

(C57BL/6N)Mouse 

Endocr

18

18 (reduced serum T4)

Immuno/ Lymphoret

19

36

36

2066

14 d1x/d(GO)

(SpragueDawley)Rat 

20

18

18

2067

36

36 (reduced in vitro production of

IgG in mitogen- stimulatedsplenocytes)

14 d1x/d(GO)

(C57BL/6N)Mouse 

21

500

500

2060

once

(C57BL/6N) (GO)Mouse 

22

36

36

2062

72

72 (reduced antibody response to

sheep red blood cells,decreased thymus weight)

14d1x/d(GO)

(C57BL/6N)Mouse 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Developmental

23

200

200

2070

10 dGd 6-15(GO)

(CD)Rat 

24

10

10

2058

25

25 (increased resorptions and

reduced fetal body weight)

10 dGd 6-15(GO)

(CD)Rat 

25

2.5

2.5

2074

10

10 (minimal increased

post-implantation loss)

10 dGd 6-15(GO)

(CD)Rat 

26

25

25

2072

50

50 (reduced fetal weight and

increased skeletal variationsassociated with maternal tox)

10 dGd 6-15(GO)

(CD)Rat 

27

1

1

b

2111

10

10 (reduced serum T4 in fetuses)

14 dGd 6-Gd 20(GO)

(LongEvans)Rat 

28

5

5 F

2045

15

15 F (delayed ossification of

sternebrae with decreasedmaternal weight gain)

13 dGd 7-19(GO)

(NewZealand)Rabbit 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

INTERMEDIATE EXPOSURESystemic

29

1.77

Hepatic 1.77 M

2039

90 d

(Sprague- (GO)Dawley)Rat 

30

14.1

Hepatic 14.1 M

2040

90 d

(Sprague- (GO)Dawley)Rat 

31

2.4

Hepatic 2.4 M

2037

90 d

(Sprague- (GO)Dawley)Rat 

32

19.2

Hepatic 19.2 M

2038

90 d

(Sprague- (GO)Dawley)Rat 

33

Hepatic

2053

9

9 (increased liver weight and

enlarged parenchymal cells)

28 d

(CD) (F)Rat 

90

Renal 90

90

Endocr 90

90

Bd Wt 90
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

34

Hepatic

2055

9

9 (increased liver weight and

enlarged parenchymal cells)

28 d

(CD) (F)Rat 

90

Renal 90

90

Endocr 90

90

Bd Wt 90
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

35

750

Resp 750 F

2077

13 wk

(CD) (F)Rat 

750

Cardio 750 F

750

Gastro 750 F

70

Hemato 70 F

750

750 F (reduced erythrocytes,

hematocrit and hemoglobin)

Hepatic

5

5 M (cytomegaly with vacuolation

and necrosis at higher doses)

50

Renal 50 M

600

600 M (minimal increase in tubular

degenerative changes)

7

Endocr 7 F

50

50 M (increased thyroid weight with

follicular epithelial changes athigher doses)

750

Dermal 750 F

750

Ocular 750 F

70

Bd Wt 70 F

600

600 M (12% reduced body weight gain)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

36

100

Resp 100

2048

90d

(Sprague- (F)Dawley)Rat 

100

Cardio 100

100

Gastro 100

100

Hemato 100

100

Musc/skel 100Hepatic

2

2 (minimal LOAEL for

hypertrophy, mild degeneration,and slight necrosis)

c

100

Renal 100

2

Endocr 2

10

10 (reduced serum T4)

100

Dermal 100

100

Ocular 100

10

Bd Wt 10

100

100 (reduced weight gain)
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

37

10

Endocr 10

2087

30

30 (reduced maternal serum T4)

36 dGd 6-Pnd 21(GO)

(LongEvans)Rat 

Immuno/ Lymphoret

38

750

750 F

2078

13 wk

(CD) (F)Rat 

39

100

100

2049

90d

(Sprague- (F)Dawley)Rat 

Reproductive

40

600

600 M

d

750

750 F

2080

13 wk

(CD) (F)Rat 

41

100

100 M

100

100 F

2051

90d

(Sprague- (F)Dawley)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

Developmental

42

1

1

a The number corresponds to entries in Figure 5- 3.b Used to derive an acute-duration oral minimal risk level (MRL) of 0.03 mg/kg/day. The MRL was derived by dividing the NOAEL by an uncertainty factor of 3from animals to humans, 3 for human variability because effects were observed in a sensitive subgroup).c Used to derive an intermediate-duration oral minimal risk level (MRL) of 0.007 mg/kg/day. The MRL was derived by dividing the LOAEL by an uncertainty faconverting a minimal LOAEL to an NOAEL, 10 for extrapolation from animals to humans, 10 for human variability)d Differences in levels of health effects and cancer effects between male and females are not indicated in Figure 5-3. Where such differences exist, only the lemost sensitive gender are presented.Bd Wt = body weight; Cardio = cardiovascular; d = day(s); (DW) = drinking water; endocr = endocrine; (F)= feed; F = female; Gd = gestation day; (G) = gavaggastrointestinal; (GO) = gavage in oil; hemato = hematological; hr = hour(s); LD50 = lethal dose; 50% kill; LOAEL = lowest observed adverse effect level;min = Musc/skel = musculoskeletal; NOAEL = no observed adverse effect level; NS = not specified; pnd = post natal day; Resp = respiratory; T4 = thyroxine ; wk = w

2086

10

10 (reduced serum T4 in fetuses

and offspring on Gd 20 and Pnd4 and 14)

36 dGd 6-Pnd 21(GO)

(LongEvans)Rat 
0.01

0.1

1

10

100

1000

10000

Death

1r 2r 3r

Hepatic

16m

17m18m18m

5r6r8r

Endocrine

15m

16m16m

17m 18m 7r 8r

13r

13r

14r

14r

Body Weight

17m

4r

4r9r10r

11r

12r

12r

Immuno/Lymphor

20m

20m

21m22m

22m 19r

mg/kg/day

Figure 5-3. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- Oral

Acute (≤14 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic
0.001

0.01

0.1

1

10

100

1000

Respiratory

35r36r

Cardiovascular

35r36r

Gastrointestinal

35r36r

Hematological

35r35r36r

Musculoskeletal

36r

Hepatic

29r30r31r32r33r 34r

35r36r

Renal

33r 34r

35r35r

36r

E

33r 

35r35r36r 

mg/kg/day

Figure 5-3. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- Oral (Cont

Intermediate (15-364 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic
0.001

0.01

0.1

1

10

100

1000

Developmental

42r42r

mg/kg/day

Figure 5-3. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- Oral (Continu

Intermediate (15-364 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Dermal

35r36r

Ocular

35r36r

Body Weight

33r 34r

35r35r36r

36r

Immuno/Lymphor

38r39r

Reproductive

40r41r

Systemic
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

ACUTE EXPOSURESystemic

1

95.9

Hepatic 95.9 M

2041

14 d1x/d(GO)

(SpragueDawley)Rat 

2

5000

Bd Wt 5000

2085

once

Spartan (GO)Rat 

3

16000

Bd Wt 16000

2002

14d1x/d(F)

(Fischer- 344)Rat 

4

100

Endocr 100

2083

4 d1x/d(GO)

(LongEvans)Rat 

5

19000

Bd Wt 19000

2005

14d1x/d(F)

(B6C3F1)Mouse 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

INTERMEDIATE EXPOSURESystemic

6

90

Hepatic 90

2057

28 d

(CD) (F)Rat 

90

Renal 90

90

Endocr 90

90

Bd Wt 90
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

7

8000

Resp 8000

2008

13 wk

(Fischer- 344) (F)Rat 

8000

Cardio 8000

8000

Gastro 8000

8000

Hemato 8000

8000

Musc/skel 8000

8000

Hepatic 8000

8000

Renal 8000

8000

Endocr 8000

8000

Bd Wt 8000
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

8

9500

Resp 9500

2013

13 wk

(B6C3F1) (F)Mouse 

9500

Cardio 9500

9500

Gastro 9500

9500

Hemato 9500

9500

Musc/skel 9500

9500

Hepatic 9500

9500

Renal 9500

9500

Endocr 9500

9500

Bd Wt 9500

Immuno/ Lymphoret

9

8000

8000

2009

13 wk

(Fischer- 344) (F)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

10

9500

9500

2014

13 wk

(B6C3F1) (F)Mouse 

Reproductive

11

8000

8000

2011

13 wk

(Fischer- 344) (F)Rat 

12

9500

9500

2016

13 wk

(B6C3F1) (F)Mouse 

Developmental

13

1000

1000 F

b

2112

20 dGd 0-19(GO)

(SpragueDawley)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

CHRONIC EXPOSURESystemic

14

1

Resp 1

2030

2 yr

(Sprague- (F)Dawley)Rat 

1

Cardio 1

1

Gastro 1

1

Hemato 1

1

Musc/skel 1

1

Hepatic 1

1

Renal 1

1

Endocr 1

1

Ocular 1

1

Bd Wt 1
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

15

2550

Resp 2550 F

2018

103 wk

(Fischer- 344) (F)Rat 

2550

Cardio 2550 F

1120

Gastro 1120 M

2240

2240 M

2550

Hemato 2550 F

2550

Musc/skel 2550 F

1200

Hepatic 1200 M

1120

1120 M

2250

2250 F (pre-neoplastic nodules)

2550

Renal 2550 F

2550

Endocr 2550 F

2550

Bd Wt 2550 F
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

16

7780

Resp 7780 F

2024

103 wk

(B6C3F1) (F)Mouse 

7780

Cardio 7780 F

3760

Gastro 3760 F

7780

7780 F (ulcers)

7780

Hemato 7780 F

7780

Musc/skel 7780 FHepatic

3200

3200 M (centrilobular hypertrophy and

granulomas)

7780

Renal 7780 FEndocr

3200

3200 M (follicular cell hyperplasia)

7780

Bd Wt 7780 F

Immuno/ Lymphoret

17

1

1

2031

2 yr

(Sprague- (F)Dawley)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

18

2019

2240

2240 M (splenic fibrosis and lymphoid

hyperplasia)

1200

1200 F (splenic hematopoiesis)

103 wk

(Fischer- 344) (F)Rat 

19

7780

7780 F

2025

103 wk

(B6C3F1) (F)Mouse 

Reproductive

20

1

1 M

1

1 F

2033

2 yr

(Sprague- (F)Dawley)Rat 

21

2240

2240 M

2550

2550 F

2021

103 wk

(Fischer- 344) (F)Rat 
LOAEL

NOAEL Serious Less Serious a

System

Key to

figure

Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (continued)

(mg/kg/day) (mg/kg/day) (mg/kg/day) 

Exposure/

Duration/

Frequency

(Specific Route) Species (Strain)

22

6650

6650 M

7780

7780 F

2027

103 wk

(B6C3F1) (F)Mouse 

Cancer

23

2022

1120

1120 M (CEL: liver neoplastic nodules)

103 wk

(Fischer- 344) (F)Rat 

24a The number corresponds to entries in Figure 5 4.b Used to derive an intermediate-duration (15-364 days) oral minimal risk level (MRL) of 10 mg/kg/day for decaBDE. The MRL was derived by dividing the 1,0by an uncertainty factor of 100 (10 for species to species extrapolation and 10 for human variability).c Differences in levels of health effects and cancer effects between male and females are not indicated in Figure 5-4. Where such differences exist, only the lemost sensitive gender are presented.Bd Wt = body weight; Cardio = cardiovascular; CEL = cancer effect level; d = day(s); (DW) = drinking water; endocr = endocrine; (F)= feed; F = female; Gd = ggavage; gastro = gastrointestinal; (GO) = gavage in oil; hemato = hematological; hr = hour(s); LOAEL = lowest observed adverse effect level; M = male; min = musculoskeletal; NOAEL = no observed adverse effect level; NS = not specified; Resp = respiratory; wk = week(s); x = time(s); yr = year(s)

2028

3200

3200 M (CEL: hepatocellular adenomas

and carcinomas)

103 wk

(B6C3F1) (F)Mouse 
10

100

1000

10000

100000

Hepatic

1r

Endocrine

4r

Body Weight

5m2r3r

mg/kg/day

Figure 5-4. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral

Acute (≤14 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic
10

100

1000

10000

Respiratory

8m7r

Cardiovascular

8m7r

Gastrointestinal

8m7r

Hematological

8m7r

Musculoskeletal

8m7r

Hepatic

8m6r7r

Renal

8m6r7r

Endocrine

8m6r7r

Body Weight

8m6r7r

Immuno/Lymphor

10m9r

R

11

mg/kg/day

Figure 5-4. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (Continue

Intermediate (15-364 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic
0.1

1

10

100

1000

10000

Respiratory

16m14r15r

Cardiovascular

16m14r15r

Gastrointestinal

16m16m14r15r15r

Hematological

16m14r15r

Musculoskeletal

16m14r15r

Hepatic

16m

14r15r

Renal

16m14r15r

Endocrine

16m14r15r

Ocular

14r

Body Weight

16m14r15r

Immu

19m17r18r

mg/kg/day

Figure 5-4. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (Continue

Chronic (≥365 days)

c-Catd-Dog r-Ratp-Pig q-Cow

 -Humansk-Monkey m-Mouseh-Rabbita-Sheep

f-Ferretj-Pigeon e-Gerbils-Hamsterg-Guinea Pig

n-Minko-Other

 Cancer Effect Level-Animals LOAEL, More Serious-Animals LOAEL, Less Serious-Animals NOAEL - Animals

 Cancer Effect Level-Humans LOAEL, More Serious-Humans LOAEL, Less Serious-Humans NOAEL - Humans

Systemic

*Doses represent the lowest dose tested per study that produced a tumorigenicresponse and do not imply the existence of a threshold for the cancer endpoint.
PBBs and PBDEs 163 

5. HEALTH EFFECTS 

general statement indicated that the rats had hunchback posture, rough coat, and sunken eyes, were 

lethargic, and appeared dehydrated and emaciated. No deaths occurred in rats administered octabromobiphenyl mixture in a single dose ≤1,000 mg/kg with 4 weeks of observation (Lee et al. 1975a), 

2,000 mg/kg with 2 weeks of observation (Norris et al. 1975a), 17,000 mg/kg with 1 week of observation 

(Lee et al. 1975a; Waritz et al. 1977), or 3,000 mg/kg/day on 2 consecutive days with 4 weeks of 

observation (Lee et al. 1975a). The 17,000 mg/kg dose was the highest that was feasible to administer, 

apparently due to gavage volume because it had to be administered as divided doses given in a 4-hour 

period. Dietary administration of octabromobiphenyl mixture in estimated dosages of ≤70 mg/kg/day for 

2 weeks was not lethal in rats, but there was no posttreatment observation period (Lee et al. 1975b; Waritz 

et al. 1977). In the only study of a decabromobiphenyl mixture, a single dose as high as 5,000 mg/kg 

caused no deaths in rats observed for 14 days (Millischer et al. 1980). In mice, dietary administration of 

FireMaster BP-6 for 2 weeks produced death (cause not reported) at estimated doses of 130 mg/kg/day, 

but not ≤36 mg/kg/day (Cagen et al. 1977; Fraker 1980; Fraker and Aust 1978). Information on acute 

oral lethality in species other than rats and mice was not located. 

In intermediate-duration studies with rats, no deaths were induced by dietary administration of FireMaster 

BP-6 at estimated dosages of ≤5 mg/kg/day for ≤82 days (Darjono et al. 1983) or ≤10 mg/kg/day for 

30 days (Akoso et al. 1982a). No deaths were observed in rats fed ≤50 mg/kg/day of an unspecified PBB 

mixture for 30 or 60 days (Sleight and Sanger 1976). Twice weekly gavage with 100 mg/kg FireMaster 

FF-1 in corn oil for two 3-week dosing periods, separated by ≈6 weeks, was not lethal in rats observed for 

2 years (Kimbrough et al. 1981). Twenty-two gavage doses of 100 mg/kg FireMaster FF-1 in corn oil 

(5 days/week for 4.5 weeks) produced 38 and 100% mortality in male and female rats, respectively; the 

average times to death were 46.7 and 60.3 days, respectively (Gupta and Moore 1979). Similar treatment 

with 30 mg/kg/day FireMaster FF-1 was not lethal in rats observed for ≈5 months. Based on these gavage 

data, the calculated LD50 in rats observed for ≈60 days posttreatment (i.e., 90-day LD50) was 149 and 

65 mg/kg/day for male and female rats, respectively (Gupta and Moore 1979). This study did not 

specifically address the cause of death, but emaciated appearance and gross loss of subcutaneous and 

visceral adipose tissue indicate wasting was a contributing factor. Rats that were treated with FireMaster 

FF-1 in corn oil by gavage on 5 days/week for 25 weeks exhibited dose-related decreased survival at 

≥0.3 mg/kg/day (cause of death not discussed), but not at 0.1 mg/kg/day (NTP 1983). The decreased 

survival was only apparent when the rats were observed for a lifetime (≈15–22 months posttreatment) and 

consistent only in males. Survival was also decreased in male but not female rats given ≥0.5 mg/kg/day 

FireMaster FF-1 in the diet for up to 104 weeks (Chhabra et al. 1993; NTP 1992). The decreased survival 

appeared to be related to increased incidences of mononuclear cell leukemia. No deaths were observed in 
PBBs and PBDEs 164 

5. HEALTH EFFECTS 

rats treated with octabromobiphenyl mixture in the diet at estimated dosages of ≤71 mg/kg/day for 

4 weeks and observed for ≤18 weeks (Lee et al. 1975b; Waritz et al. 1977). Rats treated with 

≤1 mg/kg/day dietary octabromobiphenyl mixture for 8 months did not die, but there were some deaths 

(number and cause not reported) in rats treated with higher dietary dosages (8–800 mg/kg/day) for 

30 days (Norris et al. 1975a). Insufficient information is available to determine if the deaths were 

treatment-related, since incidences and other pertinent information were not reported. 

Survival data for intermediate-duration exposure to PBBs are less extensive for species other than rat, but 

indicate that guinea pigs and mink are particularly susceptible. High mortality occurred in guinea pigs 

fed estimated dosages of 2 mg/kg/day FireMaster BP-6 for 45 days (Vos and van Genderen 1973, 1974) 

or ≥4 mg/kg/day of an unspecified PBB mixture for 30 days (Sleight and Sanger 1976); dosages of 

≤0.4 mg/kg/day of either mixture were not lethal. The Litchfield and Wilcoxon procedure was used to 

calculate dietary LD50 values of 0.47 and 0.61 mg/kg/day (estimated dosages) for male and female mink, 

respectively, exposed to FireMaster FF-1 for life (63–294 days) (Aulerich and Ringer 1979; Ringer et al. 

1981). Dosages ≤0.18 mg/kg/day did not significantly increase mortality in the mink. Dietary 

administration of FireMaster BP-6 in an estimated dosage of 21.7 mg/kg/day for 12 weeks caused some 

deaths in mice (number not reported), leading to sacrifice of other test animals (Martino et al. 1981; 

Wilson-Martino et al. 1980). Mean survival time decreased significantly in female mice treated with 

10 mg/kg/day of FireMaster FF-1 in corn oil by gavage on 5 days/week for 25 weeks, but not 

≤3 mg/kg/day (NTP 1983). Decreased survival was only apparent when the mice were observed for 

≤24 months posttreatment (lifetime observation) and not observed in similarly treated males. Survival 

was also decreased in female mice given ≥1.3 mg/kg/day FireMaster FF-1 in the diet for up to 105 weeks; 

decreased survival occurred in similarly treated male mice at 3.9 mg/kg/day (Chhabra et al. 1993; NTP 

1992). The cause of death was not discussed in the NTP (1983, 1992) mouse studies, but hepatocellular 

tumors increased significantly in both sexes at dosages that decreased survival. 

No deaths occurred in two swine that ingested estimated dosages of ≤8 mg/kg/day for 16 weeks; one pig 

was observed for 102 days following exposure and the other for 14 weeks following exposure (Ku et al. 

1978). An adult male monkey died after consuming 0.73 mg/kg/day FireMaster FF-1 in the diet for 

25 weeks (Allen et al. 1978; Lambrecht et al. 1978). The death was attributed to severe gastrointestinal 

changes, including ulcerative colitis. The only other animal in this study was a juvenile female who 

survived 50 weeks of a dietary dosage of 1.43 mg/kg/day. In another study, one juvenile female monkey 

that consumed 18 mg/kg/day FireMaster FF-1 in the diet died after 137 days of continuous exposure 

(Allen et al. 1978). Although only one or two monkeys were tested in these studies, effects characteristic 
PBBs and PBDEs 165 

5. HEALTH EFFECTS 

of PBB poisoning (e.g., dermal changes, body weight loss) indicate that the deaths were exposure-related. 

Pregnant cows given 67 mg/kg/day FireMaster BP-6 in capsules for 60 days (dosing began ≥10 days after 

pregnancy diagnosis) were sacrificed between days 33 and 66 because of impending death (Moorhead et 

al. 1977). Clinical signs developed progressively and included anorexia, emaciation, and depressed 

general condition. No mortality occurred in cows treated with ≤0.65 mg/kg/day and observed for 1 or 

140 days following the end of treatment. 

The LD50 value and reliable LOAEL values for death in each species in the acute- and intermediateduration categories are recorded in Table 5-2 and plotted in Figure 5-2. 

Polybrominated Diphenyl Ethers. No deaths occurred in rats that were treated with a single gavage dose 

of ≤5,000 mg/kg of decaBDE or ≤2,000 mg/kg of 77.4% decaBDE (containing 21.8% nonaBDE and 

0.8% octaDBE) and observed for the following 14 days (IRDC 1974; Norris et al. 1975b). No mortality 

was observed in rats and mice that were exposed to decaBDE via diet in estimated doses of ≤16,000 and 

≤19,000 mg/kg/day, respectively, for 14 days (NTP 1986). 

In intermediate-duration dietary studies with decaBDE, there was no exposure-related mortality in rats 

that were exposed to estimated dietary doses of ≤90 mg/kg/day for 28 days (IRDC 1976) or rats and mice 

fed estimated doses of ≤8,000 and ≤9,500 mg/kg/day, respectively, for 13 weeks (NTP 1986). In chronic 

studies, there were no effects on survival in rats that were fed 0.01–1.0 mg/kg/day of a 77.4% decaBDE 

mixture (containing 21.8% nonaBDE and 0.8% octaDBE) for 2 years (Kociba et al. 1975; Norris et al. 

1975b), or in rats and mice fed decaBDE in estimated doses of ≤2,550 and ≤7,780 mg/kg/day, 

respectively, for 103 weeks (NTP 1986). 

No deaths occurred in rats that were administered octaBDE by gavage in single doses ≤5,000 mg/kg and 

observed for the following 14 days (IRDC 1975a). Intermediate-duration dietary studies with octaBDE, 

observed no mortality in rats exposed to estimated dietary doses of ≤90 mg/kg/day for 28 days or 

≤750 mg/kg/day for 13 weeks (IRDC 1976, 1977). 

Single-dose gavage LD50 values of 5,000 and 6,200 mg/kg were determined for pentaBDE (Saytex 115 

and DE-71, respectively) in rats that were observed for 14 days (British Industrial Biological Research 

Association 1977; Pharmakon Research International Inc. 1984). Another study found that a single 

5,000 mg/kg dose of pentaBDE caused deaths in four of five rats in the 14 days following treatment, 

whereas doses ≤500 mg/kg caused no mortality (IRDC 1975b). No deaths occurred in rats exposed to 
PBBs and PBDEs 166 

5. HEALTH EFFECTS 

pentaBDE in estimated dietary doses of ≤90 mg/kg/day for 28 days (IRDC 1976) or ≤100 mg/kg/day for 

90 days (WIL Research Laboratories 1984). 

The LD50 and LOAEL values for death in the acute-duration BDE studies in rats are recorded in 

Tables 5-3 (lower BDEs) and 5-4 (decaBDE) and plotted in Figures 5-3 (lower BDEs) and 5-4 

(decaBDE). 

5.2.2.2 Systemic Effects 

The systemic effects in humans and animals following oral exposure to PBBs and PBDEs are described 

below. The highest NOAEL and all LOAEL values from each reliable study for systemic end points in 

each species and duration category are recorded in Tables 5-2 (PBBs), 5-3 (lower PDBEs), or 5-4 

(decaBDE) and plotted in Figures 5-2 (PBBs), 5-3 (lower PDBEs), or 5-4 (decaBDE). 

Respiratory Effects. 

Polybrominated Biphenyls. No studies were located regarding respiratory effects in humans after oral 

exposure to PBBs. 

The preponderance of data does not indicate that PBBs are respiratory system toxicants in animals, even 

at doses sufficient to cause death. No exposure-related histological changes were observed in the lungs or 

trachea of rats that were administered FireMaster FF-1 in a single dose of 200 or 1,000 mg/kg and 

observed for 2–23 months posttreatment (Kimbrough et al. 1978b, 1981). Rats and mice exposed to 

≤30 mg/kg/day FireMaster FF-1 for 2 weeks also showed no histological alterations in the lung, trachea, 

or nasal turbinates (Gupta et al. 1981). Information on acute-duration respiratory effects in other species 

was not located. 

In intermediate- and chronic-duration studies with rats, histology of the lung, trachea, or nasal turbinate 

was not altered by FireMaster FF-1 or FireMaster BP-6 dosages of ≤30 mg/kg/day by gavage for 30 days 

(Gupta et al. 1981), ≤10 mg/kg/day in the diet for 30 days (Akoso et al. 1982a; Sleight et al. 1978), 

≤10 mg/kg/day by gavage for 25 weeks (NTP 1983), or ≤1.5 mg/kg/day in the diet for up to 104 weeks 

(NTP 1992). Rat lung histology also was not affected by exposure to 50 mg/kg/day of an unspecified 

PBB mixture in the feed for 30 days (Sleight and Sanger 1976). In studies with mice, FireMaster FF-1 

produced no histopathological changes in the lungs, trachea, or nasal turbinates following gavage 
PBBs and PBDEs 167 

5. HEALTH EFFECTS 

exposure to ≤10 mg/kg/day for 25 weeks (NTP 1983) or ≤30 mg/kg/day for 30 days (Gupta et al. 1981), 

or dietary exposure to ≤3.9 mg/kg/day for up to 105 weeks (NTP 1992). Guinea pig lung histology was 

unaffected by exposures of ≤20 mg/kg/day of an unspecified PBB mixture in the feed for 30 days (Sleight 

and Sanger 1976). Relative lung weights increased in mink that died following exposure to 

≤2.4 mg/kg/day FireMaster FF-1 for 313 days, but it is unclear if this effect is adverse because the 

animals had lost weight and histopathology was not reported (Aulerich and Ringer 1979; Ringer et al. 

1981). Effects in six pregnant cows given 67 mg/kg/day of FireMaster BP-6 in capsules for up to 60 days 

included increased respiratory rate and occasional nasal discharge (Moorhead et al. 1977). Necropsy 

following sacrifice between days 33 and 66 because of impending death due to poor health showed gross 

pneumonia (one cow), microscopic lesions of early purulent bronchopneumonia (two cows), and petechial 

hemorrhages of the tracheal mucosa (one cow). No histological changes were observed in the trachea or 

lungs treated with ≤0.65 mg/kg/day and observed for 1–140 days following the end of treatment. 

Information on respiratory effects of octabromobiphenyl mixture or other PBB mixtures in animals was 

not located. 

Polybrominated Diphenyl Ethers. No studies were located regarding respiratory effects in humans after 

oral exposure to PBDEs. 

Effects of PBDEs on respiratory function have not been studied in orally-exposed animals. No 

histopathological changes in respiratory tract tissues were found in rats and mice fed decaBDE in 

estimated doses of ≤8,000 and ≤9,500 mg/kg/day, respectively, for 13 weeks (NTP 1986). In chronic 

dietary studies, there was no respiratory tract histopathology in rats that were fed ≤1.0 mg/kg/day of a 

77.4% decaBDE mixture (containing 21.8% nonaBDE and 0.8% octaDBDE) for 2 years (Kociba et al. 

1975; Norris et al. 1975b), or in rats and mice exposed to decaBDE at estimated doses of ≤2,550 and 

≤7,780 mg/kg/day, respectively, for 103 weeks (NTP 1986). 

No histopathological changes in the respiratory tract were found in dietary studies of rats exposed to 

≤750 mg/kg/day of octaBDE for 13 weeks (IRDC 1977) or ≤100 mg/kg/day of pentaBDE for 90 days 

(WIL Research Laboratories 1984). 

Cardiovascular Effects. 

Polybrominated Biphenyls. No studies were located regarding cardiovascular effects in humans after 

oral exposure to PBBs. 
PBBs and PBDEs 168 

5. HEALTH EFFECTS 

Animal data do not generally indicate cardiovascular toxicity of PBBs even at lethal doses, but 

cardiovascular function was not evaluated in most studies. No exposure-related histological changes in 

the heart were observed in rats administered FireMaster FF-1 in a single dose of 200 or ≤1,000 mg/kg and 

observed for 2–23 months posttreatment (Kimbrough et al. 1978a, 1981), or ≤30 mg/kg/day for 2 weeks 

(Gupta et al. 1981; Kimbrough et al. 1978b, 1981). Mice exposed to ≤30 mg/kg/day FireMaster FF-1 for 

2 weeks also showed no histological alterations in the heart (Gupta et al. 1981). In intermediate-and 

chronic-duration studies with rats, FireMaster FF-1 or FireMaster BP-6 dosages of ≤30 mg/kg/day by 

gavage for 30 days (Gupta et al. 1981), ≤10 mg/kg/day in the diet for 30 days (Akoso et al. 1982a; Sleight 

et al. 1978), ≤10 mg/kg/day by gavage for 25 weeks (NTP 1983), or ≤1.5 mg/kg/day in the diet for up to 

104 weeks (NTP 1992) did not alter heart weight or histology. Rat heart histology also was unaffected by 

exposure to 50 mg/kg/day of an unspecified PBB mixture for 30 days (Sleight and Sanger 1976). Rats 

exposed to ≤5 mg/kg/day FireMaster BP-6 for 30 days exhibited no exposure-related changes in blood 

pressure, but histology or other cardiovascular end points were not evaluated (McCormack et al. 1978). 

Rats exposed to octabromobiphenyl mixture in dosages of ≤1 mg/kg/day for 8 months or ≤800 mg/kg/day 

for 30 days showed no changes in heart weight, but histology or function was not evaluated (Norris et al. 

1975a). In studies with mice, FireMaster FF-1 produced no changes in heart weight or histology 

following gavage exposure to ≤10 mg/kg/day for 25 days (Gupta et al. 1981), ≤30 mg/kg/day for 30 days 

(NTP 1983), or dietary exposure to ≤3.9 mg/kg/day for up to 105 weeks (Chhabra et al. 1993; NTP 1992). 

No effects on heart relative weight or histology were reported in mink that died following exposure to 

≤2.4 mg/kg/day FireMaster FF-1 for 313 days (Aulerich and Ringer 1979; Ringer et al. 1981). Relative 

heart weights were increased in swine exposed to ≤8 mg/kg/day of an unspecified PBB mixture for 

16 weeks, but gross pathology was normal and histology was not evaluated (Ku et al. 1978). Necropsy of 

a monkey that died following ingestion of 0.73 mg/kg/day FireMaster FF-1 for 25 weeks showed an 

enlarged heart, but histology was not evaluated and a similar effect was not reported in two other 

monkeys exposed to higher dosages (Allen et al. 1978; Lambrecht et al. 1978). Mean heart rate was 32% 

lower than the pre-exposure value in pregnant cows that were treated with 67 mg/kg/day of FireMaster 

BP-6 in capsules for 10 days (Moorhead et al. 1977). Necropsy following sacrifice between days 33 and 

66 because of impending death due to poor health showed petechial and ecchymotic hemorrhages of the 

myocardium and endocardium in two of six cows. No cardiovascular effects were observed in cows 

given ≤0.65 mg/kg/day and observed for 1–140 days following the end of treatment. 

Polybrominated Diphenyl Ethers. No studies were located regarding cardiovascular effects in humans 

after oral exposure to PBDEs. 
PBBs and PBDEs 169 

5. HEALTH EFFECTS 

Effects of PBDEs on cardiovascular function have not been studied in orally exposed animals. No 

histopathological changes in the heart were found in rats and mice fed decaBDE in estimated doses of 

≤8,000 and ≤9,500 mg/kg/day, respectively, for 13 weeks (NTP 1986). In chronic dietary studies, there 

was no cardiac histopathology in rats that were fed ≤1.0 mg/kg/day of a 77.4% decaBDE mixture 

(containing 21.8% nonaBDE and 0.8% octaDBDE) for 2 years (Norris et al. 1975b), or in rats and mice 

exposed to decaBDE in estimated doses of ≤2,550 and ≤7,780 mg/kg/day, respectively, for 103 weeks 

(NTP 1986). 

No histopathological changes in the heart were found in dietary studies of rats exposed to 

≤750 mg/kg/day of octaBDE for 13 weeks (IRDC 1977) or ≤100 mg/kg/day of pentaBDE for 90 days 

(WIL Research Laboratories 1984). 

Gastrointestinal Effects. 

Polybrominated Biphenyls. No symptoms of gastrointestinal effects were reported by residents of 

quarantined Michigan farms in an epidemiological study conducted by the U.S. Center for Disease 

Control and the Michigan Department of Public Health (Landrigan et al. 1979). In a medical history 

survey conducted by the Environmental Science Laboratory of the Mount Sinai School of Medicine, no 

statistically significant difference was observed between the prevalence rates of gastrointestinal 

symptoms for 933 Michigan residents who were likely to have ingested PBB-contaminated food and the 

rates for a control group of 229 Wisconsin farm residents (Anderson et al. 1978c). The Michigan 

residents were examined ≈3 years after the contamination episode occurred. No other studies were 

located regarding gastrointestinal effects in humans after oral exposure to PBBs. 

Gastric lesions have developed in various animals that ingested PBBs, particularly after prolonged 

exposure to FireMaster FF-1 or FireMaster BP-6. No exposure-related histological changes in the 

gastrointestinal tract or esophagus were observed in rats administered FireMaster FF-1 in a single dose 

≤1,000 mg/kg and observed for 2–23 months posttreatment (Kimbrough et al. 1978b, 1981). Rats or mice 

exposed to ≤30 mg/kg/day FireMaster FF-1 for 2 weeks also showed no histological alterations in the 

gastrointestinal tract (esophagus not examined) (Gupta et al. 1981). In intermediate-duration studies, the 

gastrointestinal tract of rats exposed to FireMaster BP-6 or FireMaster FF-1 by gavage or diet at 

≤50 mg/kg/day for 4–4.5 weeks showed no histopathological changes (esophagus not examined) (Akoso 

et al. 1982a; Gupta and Moore 1979; Gupta et al. 1981; Sleight and Sanger 1976; Sleight et al. 1978). 

Histological examination of the gastrointestinal tract of rats administered FireMaster FF-1 by gavage for 

25 weeks showed significantly increased incidences of gastric ulcers at ≥1 mg/kg/day and hyperplastic 
PBBs and PBDEs 170 

5. HEALTH EFFECTS 

gastropathy at ≥3 mg/kg/day after lifetime observation (23 months). These gastric effects were not 

observed in rats examined at the end of the gavage treatment period, although similar changes 

(forestomach hyperplasia, inflammation, and ulceration) occurred in rats exposed to 1.5 mg/kg/day 

FireMaster FF-1 in the diet for up to 104 weeks (Chhabra et al. 1993; NTP 1992). In the only study of a 

decabromobiphenyl mixture, rats were fed estimated dosages as high as 100 mg/kg/day for 13 weeks 

(Millischer et al. 1980). A comprehensive histology evaluation was performed in this study, but the liver 

is the only tissue specifically mentioned as having been examined. Due to the total number of tissues 

examined (21) and route of exposure, it is probable that the gastrointestinal tract was examined but not 

discussed because no histopathologic changes were observed. 

Gastrointestinal tract histology was normal in mice exposed to FireMaster FF-1 dosages of 

≤10 mg/kg/day by gavage for 25 weeks or ≤3.9 mg/kg/day in the diet for up to 105 weeks (NTP 1983, 

1992). FireMaster FF-1 produced no histological changes in the gastrointestinal tract of mice exposed to 

≤30 mg/kg/day for 30 days (Gupta et al. 1981). Gross pathologic examination of swine administered an 

unspecified PBB mixture for 16 weeks showed that the glandular portion of the stomach "appeared 

somewhat hyperplastic" (additional details were not reported, and histology was not evaluated) at 

8 mg/kg/day, but not at 1 mg/kg/day (Ku et al. 1978). Biopsies of two monkeys performed following 

their ingestion of 0.73 or 1.43 mg/kg/day FireMaster FF-1 for 12 weeks showed proliferation of gastric 

mucosal cells, focal areas of infiltration of chronic inflammatory cells, and isolated penetrations of the 

gastric mucosa into the underlying submucosa (Allen et al. 1978; Lambrecht et al. 1978). Necropsies 

performed after 25 or 50 weeks of exposure also showed hyperplastic gastroenteritis and, in the low-dose 

monkey (that died of "severe gastrointestinal changes"), severe ulcerative colitis. Hyperplastic 

gastroenteritis was described in another monkey exposed to a higher dosage (18 mg/kg/day) of FireMaster 

FF-1 for 137 days (Allen et al. 1978). Gastrointestinal effects in six pregnant cows that were 

administered 67 mg/kg/day of FireMaster BP-6 in capsules for up to 60 days included diarrhea, 

dehydration (possibly a result of the diarrhea), and occasional constipation (Moorhead et al. 1977). 

Necropsy following sacrifice between days 33 and 66 because of impending death due to poor health 

showed edema and hemorrhage of the colon and rectum mucosa, although histology was normal in the 

esophagus, rumen, omasum, and reticulum. No histological changes were observed in the gastrointestinal 

tract of cows with ≤0.65 mg/kg/day and observed 1 or 140 days following the end of treatment. 

Polybrominated Diphenyl Ethers. No studies were located regarding gastrointestinal effects in humans 

after oral exposure to PBDEs. 
PBBs and PBDEs 171 

5. HEALTH EFFECTS 

No histopathological changes in gastrointestinal tract tissues were found in rats and mice fed decaBDE in 

estimated doses of ≤8,000 and ≤9,500 mg/kg/day, respectively, for 13 weeks (NTP 1986). In chronic 

dietary studies, there was no gastrointestinal tract histopathology in rats that were fed ≤1.0 mg/kg/day of a 

77.4% decaBDE mixture (containing 21.8% nonaBDE and 0.8% octaDBDE) for 2 years (Kociba et al. 

1975; Norris et al. 1975b). Higher dietary doses of decaBDE for 103 weeks caused acanthosis of the 

forestomach in rats exposed to 2,240 mg/kg/day (no effects at ≤1,200 mg/kg/day) and stomach ulcers in 

mice exposed to 7,780 mg/kg/day (no effects at ≤3,760 mg/kg/day) (NTP 1986). 

No histopathological changes in the gastrointestinal tract were found in dietary studies of rats exposed to 

≤750 mg/kg/day of octaBDE for 13 weeks (IRDC 1977) or ≤100 mg/kg/day of pentaBDE for 90 days 

(WIL Research Laboratories 1984). 

Hematological Effects. 

Polybrominated Biphenyls. No studies were located regarding hematological effects in humans after oral 

exposure to PBBs. 

In animals, hematologic changes indicative of possible anemia are common findings in animals resulting 

from longer-term exposure to PBBs. Comprehensive hematological examinations in rats and mice 

administered ≤30 mg/kg/day FireMaster FF-1 for 2 weeks showed no exposure-related changes (Gupta et 

al. 1981). No additional information on hematology in animals following acute-duration exposure to 

PBBs was located. In intermediate-duration studies, no consistent hematological changes were found in 

rats exposed to ≤10 mg/kg/day FireMaster BP-6 for 30 days (Akoso et al. 1982a; Sleight et al. 1978). 

Some hematologic effects occurred in rats at higher dosages or longer durations. Exposure to 

30 mg/kg/day FireMaster FF-1 for 4.5 weeks significantly reduced hemoglobin concentration, packed cell 

volume (PCV), and platelet count in rats evaluated up to ≈60 days postexposure (Gupta and Moore 1979). 

In another study in which rats were administered the same dosages of FireMaster FF-1 (≤30 mg/kg/day) 

for 30 days, longer postexposure (up to 90 days) evaluation revealed transient responses (Gupta et al. 

1981). Transitory and slight but significant (p<0.05) decreases in red blood cell count, hemoglobin 

concentration, and PCV values were found; they returned to control levels by 60-days post-dosing. No 

consistent hematological changes were observed in rats administered ≤50 mg/kg/day of an unspecified 

PBB mixture for 30 days (Sleight and Sanger 1976) or ≤10 mg/kg/day FireMaster FF-1 for 6 months 

(Luster et al. 1980). Rats exposed to FireMaster FF-1 for 25 weeks showed no hematological changes at 

0.1 mg/kg/day, but had dose-related, significantly decreased hemoglobin, mean corpuscular hemoglobin 

(MCH), mean corpuscular volume (MCV), and PCV at ≥0.3 mg/kg/day, and increased total leukocytes at 
PBBs and PBDEs 172 

5. HEALTH EFFECTS 

≥1 mg/kg/day; there were no effects on erythrocyte or platelet counts (NTP 1983). Mice similarly treated 

for 25 weeks had decreased erythrocyte count and MCV at ≥0.3 mg/kg/day and decreased platelets and 

lymphocytes at ≥1 mg/kg/day, but no hematological effects were noted at 0.1 mg/kg/day (NTP 1983). No 

hematologic alterations were found in mice exposed to FireMaster FF-1 at dosages of ≤10 mg/kg/day for 

6 months or ≤30 mg/kg/day for 30 days (Gupta et al. 1981; Luster et al. 1980). 

Hematologic evaluation of swine treated with an unspecified PBB mixture for 16 weeks showed 

significantly decreased hemoglobin and hematocrit values in two of four animals exposed to 8 mg/kg/day 

at week 6, after which values returned to normal or near-normal within 2 weeks (Ku et al. 1978). 

Decreased PCV and serum protein developed in monkeys exposed to FireMaster FF-1 in dosages of 

≥0.73 mg/kg/day for ≥25 weeks (two animals); additional hematologic effects observed in one monkey 

exposed to 18 mg/kg/day for 137 days were decreased erythrocyte and white blood cell counts (Allen et 

al. 1978; Lambrecht et al. 1978). No hematological changes were measured in cows treated with 

≤0.65 mg/kg/day FireMaster BP-6 in capsules for 60 days, and observed for up to 140 days following the 

end of treatment (Moorhead et al. 1977). Similar treatment with 67 mg/kg/day did not cause abnormal 

hematologic indices in four of six cows; changes in the other two animals (e.g., leukocytosis, increased 

PCV) have uncertain toxicologic significance because the animals at this dose were sacrificed 

between days 33 and 66 because of impending death due to poor health. 

Studies of hematologic effects of octabromobiphenyl mixture, performed only in rats, showed 

significantly decreased red blood cell count and PCV following 800 mg/kg/day for 30 days, but no 

hematological changes resulting from ≤1 mg/kg/day for 8 months (Norris et al. 1975a). In the only study 

of a decabromobiphenyl mixture, dietary administration of 100 mg/kg/day for 13 weeks caused no 

hematologic changes in rats (Millischer et al. 1980). Erythrocyte and leucocyte counts, differential 

leukocyte count, and hematocrit and hemoglobin levels were measured. 

Polybrominated Diphenyl Ethers. No studies were located regarding hematological effects in humans 

after oral exposure to PBDEs. 

OctaBDE caused red blood cell effects in an intermediate-duration dietary studies in rats in which 

erythrocytes, hematocrit, and hemoglobin were reduced following exposure to 750 mg/kg/day for 

13 weeks (no effects at ≤70 mg/kg/day) (IRDC 1977). 
PBBs and PBDEs 173 

5. HEALTH EFFECTS 

DecaBDE and pentaBDE did not induce hematological effects in animals. In dietary studies with 

decaBDE, no hematological changes were found in rats exposed to ≤800 mg/kg/day for 30 days (Norris et 

al. 1973, 1975a, 1975b), ≤8,000 mg/kg/day for 13 weeks (NTP 1986), or ≤2,550 mg/kg/day for 

103 weeks (NTP 1986), or in mice exposed to ≤9,500 mg/kg/day for 13 weeks or ≤7,780 mg/kg/day for 

103 weeks (NTP 1986). There also were no hematological effects in rats exposed by diet to 

≤1.0 mg/kg/day of a 77.4% decaBDE mixture (containing 21.8% nonaBDE and 0.8% octaDBE) for 

2 years (Kociba et al. 1975; Norris et al. 1975b) or ≤100 mg/kg/day of pentaBDE for 90 days (WIL 

Research Laboratories 1984). 

Musculoskeletal Effects. 

Polybrominated Biphenyls. Symptoms of musculoskeletal effects, described as "joint pain" and 

"swelling in joints," were frequently cited health complaints in two epidemiological studies of groups of 

Michigan residents who were likely to have ingested PBB-contaminated food (Anderson et al. 1978c; 

Landrigan et al. 1979). Although one study demonstrated a statistically significant difference between the 

prevalence rate for these types of symptoms in Michigan residents compared with nonexposed residents 

of Wisconsin farms (Anderson et al. 1978c), neither study demonstrated a positive association between 

serum PBB levels and the prevalence rates for symptoms of musculoskeletal effects. 

There are no pathology data indicating that PBBs produce effects in musculoskeletal tissues of animals. 

No exposure-related histological changes in muscle or bone marrow were observed in rats that were 

administered a single 1,000 mg/kg dose of FireMaster FF-1 and observed for 2 years (Kimbrough et al. 

1981). Rats and mice exposed to ≤30 mg/kg/day FireMaster FF-1 for 2 weeks showed no histological 

alterations in muscle or sternum (Gupta et al. 1981). In intermediate-and chronic-duration studies, rats 

and/or mice exposed to FireMaster FF-1 or FireMaster BP-6 dosages of ≤30 mg/kg/day by gavage for 

30 days (Gupta et al. 1981), ≤10 mg/kg/day in the diet for 30 days (Akoso et al. 1982a; Sleight et al. 

1978), ≤10 mg/kg/day by gavage for 25 weeks (NTP 1983), or ≤3.9 mg/kg/day in the diet for up to 

105 weeks (NTP 1992) showed no histopathological changes in muscle or bone. A dosage of an 

unspecified PBB mixture as high as 50 mg/kg/day for 30 days produced no histopathological changes in 

rat muscle (Sleight and Sanger 1976). Excess porphyrins were detected in bone and/or teeth by 

fluorescence under ultraviolet light in some of the rat studies (Gupta and Moore 1979; NTP 1983), but 

this appears to be a consequence of altered porphyrin metabolism (Hill 1985). No histological alterations 

were observed in sternebrae bone marrow of pregnant cows given FireMaster BP-6 in capsules for up to 

60 days (Moorhead et al. 1977). Cows treated with 67 mg/kg/day were necropsied following sacrifice 
PBBs and PBDEs 174 

5. HEALTH EFFECTS 

between days 33 and 66 because of impending death due to poor health, and those treated with nonlethal 

lower dosages of ≤0.65 mg/kg/day were examined 1 or 140 days following the end of treatment. 

Polybrominated Diphenyl Ethers. No studies were located regarding musculoskeletal effects in humans 

after oral exposure to PBDEs. 

Dietary studies with decaBDE found no histopathological changes in musculoskeletal tissues in rats 

exposed to ≤8,000 mg/kg/day for 13 weeks (NTP 1986), ≤1.0 mg/kg/day (77.4% containing 21.8% 

nonaBDE and 0.8% octaDBE) for 2 years (Kociba et al. 1975; Norris et al. 1975b), or ≤2,550 mg/kg/day 

for 103 weeks (NTP 1986), or in mice exposed to ≤9,500 mg/kg/day for 13 weeks or ≤7,780 mg/kg/day 

for 103 weeks (NTP 1986). A study of pentaBDE found no musculoskeletal changes in rats exposed 

dietary doses of ≤100 mg/kg/day for 90 days (WIL Research Laboratories 1984). No information was 

located on possible musculoskeletal effects of octaBDE. 

Hepatic Effects. 

Polybrominated Biphenyls. Results from several studies of humans exposed to PBBs do not 

demonstrate, in general, a conclusive association between adverse effects on the liver and oral exposure to 

PBBs. In a study in which serum was collected in 1974, 1977, 1978, and 1979 from 89, 240, 220, and 

200 individuals, respectively, who were predominately residents of quarantined Michigan farms, no 

consistent statistically significant correlations were found between serum PBB levels and levels of serum 

glutamic oxaloacetic transaminase (SGOT), serum glutamic pyruvic transaminase (SGPT) or serum 

bilirubin (Kreiss et al. 1982). The prevalence rates of Michigan residents with abnormally high levels of 

SGPT (12.7% prevalence rate), SGOT (12.7%), or lactate dehydrogenase (8.6%) were statistically 

significantly higher than comparable rates for residents of Wisconsin farms (2.7, 2.0, and 3.3%) 

(Anderson et al. 1979). A contingency table analysis indicated that the prevalence of abnormal SGPT 

values in Michigan residents with serum PBB levels ≤1 ppb (8%) was lower than the prevalence rate for 

residents with serum PBB levels ≥1 ppb (14%), but correlation coefficients for serum PBB levels and 

serum liver enzyme levels were uniformly low (r<0.1) (Anderson et al. 1979). Physical examinations of 

Michigan residents (37 men and 9 women) with known exposure to PBBs and a history of incapacitating 

health care complaints revealed that 72% of the subjects displayed mildly enlarged livers, which were 

associated with elevations in serum liver enzymes (SGOT and SGPT) predominately less than 2-fold 

above normal values (Stross et al. 1979). Mildly enlarged livers, confirmed by liver scanning, were 

observed in 4 of 23 (17%) Michigan residents with known PBB exposure and incapacitating health 

complaints and in 2 of 28 (7%) workers involved in the manufacture and distribution of PBBs, 
PBBs and PBDEs 175 

5. HEALTH EFFECTS 

respectively; however, these workers had histories of either substantial alcohol intake or exposure to 

multiple chemicals (Stross et al. 1981). Results of a caffeine breath test, discussed in Section 5.8.2, 

suggest that PBBs may have induced hepatic microsomal enzymes in exposed Michigan residents 

(Lambert et al. 1990). 

Hepatic effects of PBBs are documented in various animal species although rats have been the species 

tested most extensively. The changes appear to be similar among species and reversible when mild. 

Characteristic hepatic effects include proliferation of the smooth endoplasmic reticulum, microsomal 

enzyme induction, increased serum levels of liver-associated enzymes indicative of possible 

hepatocellular damage, liver enlargement, hepatocyte vacuolation and fat deposition, fibrosis, and 

necrosis. PBBs also cause alterations in levels of cholesterol and other lipids in liver and serum, levels of 

vitamin A in liver and urine, and levels of porphyrins in liver, other tissues, and urine. These changes 

could be secondary to liver damage or due to direct effects on lipid, vitamin A, and porphyrin 

metabolism, which occurs primarily in the liver. Induction of microsomal enzymes by PBBs is a 

sensitive effect generally regarded as an adaptive response of the liver rather than as a manifestation of 

hepatotoxicity per se (Guzelian 1985). Although not necessarily adverse, induction of microsomal 

enzymes could alter the rate or pathways of metabolism of other xenobiotic or endogenous substances and 

increase activation of promutagens and procarcinogens or increase detoxification pathways. In addition, 

the induction of some microsomal enzyme activities is an indicator of exposure to PBBs and related 

compounds (AhR agonists), which elicit a well known pattern of toxic responses (see Chapters 3 and 4). 

PBB-related liver enlargement is usually associated with hepatocyte enlargement and an increase in 

smooth endoplasmic reticulum and/or increased microsomal enzymatic activity; therefore, it is not 

considered an adverse effect unless accompanied by other biochemical changes and/or histological 

alterations. 

Rats administered FireMaster FF-1 in a single 1,000 mg/kg dose and observed for 2–23 months 

posttreatment or a lethal dose of 1,000 mg/kg/day for 2 weeks developed enlarged livers with fatty and 

necrotic changes leading to fibrosis (Gupta and Moore 1979; Kimbrough et al. 1978b, 1981). Lower 

single doses of FireMaster FF-1 caused vacuolation and some biochemical changes (e.g., increased serum 

cholesterol and phospholipids, decreased serum retinol) at 500 mg/kg (Bernert et al. 1983; Kimbrough et 

al. 1980), and hepatic porphyrin accumulation with no histologic changes at 200 mg/kg (Kimbrough et al. 

1981). Repeated exposure to lower dosages of ≥3 mg/kg/day FireMaster FF-1 for 2 weeks (Gupta et al. 

1981) or 5 mg/kg/day FireMaster BP-6 for 10 days (Raber and Carter 1986) caused hepatocyte 

enlargement and some fatty and single-cell necrotic changes in weanling and young rats. A limited 
PBBs and PBDEs 176 

5. HEALTH EFFECTS 

amount of data suggest that octabromobiphenyl mixture-induced hepatic effects in rats are milder than for 

FireMaster mixtures at similar dosages. Fatty changes appear to be the most severe hepatic 

histopathologic effect of octabromobiphenyl observed following a single 1,000 mg/kg dose or doses of 

3,000 mg/kg/day for 2 days and 6.53 mg/kg/day (but not 0.66 mg/kg/day) for 2 weeks (Lee et al. 1975a, 

1975b; Waritz et al. 1977). In studies with mice, a single dose of 36 mg/kg FireMaster BP-6 increased 

liver weight (histology not evaluated) and had no consistent effects on disposition of injected ouabain or 

indocyanine green, indicating that hepatic function was not compromised (Cagen et al. 1977). Sporadic 

increases in the clearance of ouabain and indocyanine green were attributed to increased liver size. 

Exposure to 130 mg/kg/day FireMaster BP-6, for 11 days caused focal areas of coagulative necrosis 

(Corbett et al. 1975) and ≥3 mg/kg/day FireMaster FF-1 for 2 weeks caused scattered necrosis in mice 

(Gupta et al. 1981). 

In intermediate-duration studies with rats, dosages ≥0.05 mg/kg/day FireMaster BP-6 for 20 days induced 

hepatic microsomal enzymes but histology was not evaluated (Babish et al. 1978). Dose-related 

hepatocyte swelling and vacuolation were induced by ≥0.1 mg/kg/day FireMaster BP-6 for 30 days 

(Akoso et al. 1982a), lipid accumulation, porphyrin levels, and atypical foci were increased by 

≥0.3 mg/kg/day FireMaster FF-1 for 25 weeks (lethal dose) (NTP 1983), and bile duct hyperplasia was 

induced by 0.5 mg/kg/day FireMaster BP-6 for 82 days (Darjono et al. 1983). Rats exposed to higher, but 

not necessarily lethal, dosages of FireMaster FF-1 or FireMaster BP-6 for 1–3 months showed 

progression of these effects, including marked degenerative changes and porphyrin accumulation in these 

and other studies (Gupta and Moore 1979; Gupta et al. 1981; Kasza et al. 1978a; McCormack et al. 1978; 

Sleight and Sanger 1976; Sleight et al. 1978). In the only chronic study, incidences of hepatocellular 

hypertrophy, cytoplasmic vacuolation, atypical foci, and oval cell hyperplasia were increased in rats fed 

≥0.5 mg/kg/day FireMaster FF-1 for up to 104 weeks (Chhabra et al. 1993; NTP 1992). Compared to this 

adult-only exposure, combined perinatal and adult exposure resulted in increased incidences of oval cell 

hyperplasia at 0.5 mg/kg/day and hypertrophy, cytoplasmic vacuolation, and bile duct fibrosis at 

1.5 mg/kg/day. Studies of octabromobiphenyl mixture in rats have shown hepatic effects (e.g., 

hypertrophy and hyperplasia of centrilobular cells, vacuolation, and other fatty degenerative changes) at 

dosages ≥6.53 mg/kg/day for 4 weeks (Lee et al. 1975b; Norris et al. 1975a; Waritz et al. 1977), but 

normal liver histology at ≤1 mg/kg/day for 8 months (Norris et al. 1975a). A 13-week dietary study with 

decabromobiphenyl mixture found that hepatic effects in rats, including vacuolization and distension of 

centrilobular hepatocytes often accompanied by slightly increased lipid, did not occur at dosages 

<100 mg/kg/day (Millischer et al. 1980). Information on hepatic effects of octabromobiphenyl mixture 

and decabromobiphenyl in species other than the rat was not located. 
PBBs and PBDEs 177 

5. HEALTH EFFECTS 

In mice, exposure to FireMaster FF-1 for 25 weeks increased liver weight, porphyrin content, and SGOT 

at ≥0.3 mg/kg/day and hepatocyte swelling occurred at ≥1 mg/kg/day (NTP 1983). Hepatic effects in 

mice exposed to ≥1.3 mg/kg/day FireMaster FF-1 for up to 105 weeks included hepatocyte hypertrophy, 

vacuolization, and necrosis; bile duct hyperplasia also developed (NTP 1992). Dosages ≥3 mg/kg/day for 

4–6 weeks, but not 0.3 mg/kg/day, also induced hepatocyte necrosis and/or vacuolation in mice (Gupta et 

al. 1981; Loose et al. 1981; NTP 1983). Fatty changes and centrilobular necrosis developed in pregnant 

swine fed ≥1.25 mg/kg/day, but not 0.125 mg/kg/day, FireMaster BP-6 for 12 weeks during the second 

half of gestation through lactation (Werner and Sleight 1981). This adverse effect level cannot be 

corroborated in nonpregnant swine exposed to ≤8 mg/kg/day of unspecified PBBs for 16 weeks due to 

lack of liver histology evaluations, although relative liver weight increased at ≥1 mg/kg/day and no gross 

changes were observed (Ku et al. 1978). Guinea pigs appear to be particularly susceptible to hepatic 

effects of PBBs (unspecified) as indicated by ultrastructural vacuolation and formation of myelin bodies 

in hepatocytes following exposure to ≥0.04 mg/kg/day for 30 days; liver weights were increased at 

0.4 mg/kg/day and histological vacuolation and severe centrilobular fatty change were observed at a lethal 

dose of 4 mg/kg/day (Sleight and Sanger 1976). Mink that ingested ≥0.24 mg/kg/day FireMaster FF-1 for 

≤313 days showed increased liver weight and fatty infiltration (Aulerich and Ringer 1979; Ringer et al. 

1981). In monkeys, lethal FireMaster FF-1 dosages ≥0.73 mg/kg/day for 25–50 weeks caused hepatocyte 

enlargement with increased lipid droplets, bile duct hyperplasia, increased SGPT, and decreased serum 

cholesterol (Allen et al. 1978; Lambrecht et al. 1978). Effects in six pregnant cows given 67 mg/kg/day 

of FireMaster BP-6 in capsules for up to 60 days included increased serum lactic dehydrogenase (LDH) 

and SGOT (Moorhead et al. 1977). Necropsy following sacrifice between days 33 and 66 because of 

impending death due to poor health showed increased liver weight and pathologic liver changes including 

friable appearance, glycogen depletion in hepatocytes, sinusoidal dilation, and scattered areas of early 

fatty degeneration. In general, the hepatic effects observed in cows are less pronounced than in other 

species at lethal doses. No adverse hepatic effects were observed in cows treated with ≤0.65 mg/kg/day 

and examined 1 or 140 days following the end of treatment. 

Polybrominated Diphenyl Ethers. No studies were located regarding hepatic effects in humans after oral 

exposure to PBDEs. 

No information was located on adverse hepatic effects of acute-duration oral exposure to PBDEs in 

animals. Intermediate- and chronic-duration studies in rodents indicate that the liver is a target of toxicity 

of PBDEs. As discussed below, repeated dietary exposure to commercial PBDE products typically 
PBBs and PBDEs 178 

5. HEALTH EFFECTS 

caused liver enlargement with or without degenerative changes, and effects were generally dose-related in 

incidence and severity, more frequent and pronounced in males than females, and more severe and 

induced by lower doses with octaBDE and pentaBDE than decaBDE. Regarding decaBDE, the low 

purity former mixture (77.4% pure containing 21.8% nonaBDE and 0.8% octaBDE) was much more 

hepatotoxic than the current day commercial product (≈99% pure). In particular, hepatic effects included 

increased liver weights at ≥80 mg/kg/day with centrilobular cytoplasmic enlargement and vacuolation at 

800 mg/kg/day in male rats exposed to 77% decaBDE for 30 days (Norris et al. 1973, 1975a, 1975b), 

whereas exposure to 94–97% decaBDE for 13 weeks caused no liver pathology in rats and mice exposed 

to estimated doses as high as 2,000–8,000 and 2,375–9,500 mg/kg/day, respectively (NTP 1986). The 

NOAEL for hepatic effects from the NTP (1986) 13-week study (8,000 mg/kg/day) was used as the basis 

for the intermediate-duration MRL for oral exposure to decaBDE, as indicated in a footnote to Table 5-4 

and discussed in Chapter 4 and Appendix A. In chronic studies, exposure to 94–97% decaBDE for 

103 weeks caused liver lesions that included neoplastic nodules in rats at ≥1,120 mg/kg/day, thrombosis 

and degeneration in rats at 2,240 mg/kg/day, and centrilobular hypertrophy and granulomas in mice at 

≥3,200 mg/kg/day (NTP 1986). The thrombosis in the rats was characterized by a near total occlusion of 

a major heptatic blood vessel by a dense fibrin coagulum. A NOAEL was not identified in the rats or 

mice. The only other chronic study of decaBDE found that exposure to 1 mg/kg/day of the 77% pure 

mixture for 2 years caused no liver effects in rats; higher doses were not tested, precluding identification 

of a LOAEL (Kociba et al. 1975; Norris et al. 1975b). Unlike the decaBDE used in the NTP (1986) 

study, the composition of the mixture tested by Kociba et al. (1975) is substantially different than that of 

current-day commercial decaBDE products (≈99%). 

OctaBDE caused increased liver weight and histopathological changes such as hepatocellular enlargement 

and vacuolation in rats exposed to doses as low as 5mg/kg/day (lowest tested dose) for 4–13 weeks 

(IRDC 1976, 1977). Hepatic effects in rats exposed to octaBDE for 13 weeks included increases in 

absolute and/or relative liver weight at ≥5–7 mg/kg/day and liver lesions in 40% of males at 5 mg/kg/day 

and 100% of both sexes at ≥50 mg/kg/day (IRDC 1977). The lesions were dose-related in severity as 

well as incidence and characterized by cytomegaly, change in hepatocytic cytoplasm to a finely granular, 

homogeneous type, and cytoplasmic vacuolation. At ≥600 mg/kg/day many of the livers had vacuolation 

of centrolobular hepatocytes and some had hepatocyte necrosis. Examinations performed at 8 weeks and 

6 months postexposure showed that the liver effects persisted in the rats exposed to ≥50 mg/kg/day for 

13 weeks (IRDC 1977). 
PBBs and PBDEs 179 

5. HEALTH EFFECTS 

PentaBDE induced liver effects in rats exposed to ≥9 mg/kg/day for 28 days (increased liver weight and 

enlargement of centrolobular and midzonal liver parenchymal cells) (IRDC 1976) and ≥2 mg/kg/day for 

90 days (hepatocytomegaly) (WIL Research Laboratories 1984). The hepatomegaly in the 90-day study 

was dose-related with respect to severity (some affected hepatocytes at higher doses had vacuoles that 

likely contained lipid) and not completely reversible, as it was still evident in ≥10 mg/kg/day males and 

100 mg/kg/day females at 24 weeks postexposure in lessened severity and incidence. Females exposed to 

2 and 100 mg/kg/day pentaBDE for 90 days had an increased incidence of degeneration and necrosis of 

individual liver parenchymal cells at 24 weeks postexposure; the investigators concluded that this may 

represent the final loss of previously damaged cells and probably should be considered compound-related 

(WIL Research Laboratories 1984). The 2 mg/kg/day LOAEL for hepatic effects of pentaBDE was used 

as the basis for the intermediate-duration MRL for oral exposure to tetra- to heptaBDEs as indicated in 

the footnote to Table 5-3 and discussed in Chapter 4 and Appendix A. Liver vitamin A concentrations 

were increased in rats and mice exposed to a commercial pentaBDE mixture (Bromkal 70-5 DE) by 

gavage in doses of 18 and 36 mg/kg/day, respectively, for 90 days (Hallgren et al. 2001). 

Hepatic microsomal enzyme induction is a well-documented effect for the pentaBDE and octaBDE 

commercial products, but has not been observed for decaBDE (Carlson 1980b; Zhou et al. 2001). 

Microsomal enzyme activity was induced in rats exposed by gavage to doses as low as 0.6 mg/kg/day of 

octaBDE and 0.4 mg/kg/day of pentaBDE for 90 days as indicated by increases in O-ethyl O-p-nitrophenyl phenylphosphonothioate (EPN) detoxification, p-nitroanisole demethylation, and cytochrome c 

reductase and cytochrome P-450 levels (Carlson 1980a). Some of these changes were persistent, lasting 

for 30-60 days after cessation of treatment, but not considered to be adverse due to the lack of any 

accompanying hepatic histological abnormalities. Rats that were treated with equimolar 

(0.1 mmol/kg/day) gavage doses of deca-, octa-, or pentaBDE (95.9, 76.6, or 56.4 mg/kg/day, 

respectively) for 14 days had octa- and pentaBDE-induced increases in liver weight and microsomal 

enzyme activity (e.g., increased EPN detoxification, p-nitroanisole demethylation, uridinediphosphateglucuronyltransferase [UDPGT] activity, and benzo[a]pyrene hydroxylase activity); exposure to decaBDE 

only increased liver weight (Carlson 1980b). DecaBDE also had no effect on hepatic UDPGT, 

ethoxyresorufin-o-deethylase (EROD), or pentoxyresorufin-o-deethylase (PROD) activities in weanling 

rats that were treated with ≤100 mg/kg/day by gavage for 4 days (Zhou et al. 2001). Similar exposure to 

0.3–300 mg/kg/day pentaBDE caused significantly increased EROD and PROD at ≥10 mg/kg/day and 

UDPGT at ≥30 mg/kg/day, and octaBDE induced PROD at ≥10 mg/kg/day and EROD and UDPGT at 

≥30 mg/kg/day; neither of these PBDEs caused induction at ≤1 mg/kg/day. 
PBBs and PBDEs 180 

5. HEALTH EFFECTS 

Hepatic EROD, PROD, and methoxyresorufin-o-deethylase (MROD) were increased in mice exposed to 

≥18 mg/kg/day (lowest tested dose) of pentaBDE by gavage for 14 days, although UDPGT was 

unchanged at ≤36 mg/kg/day (Fowles et al. 1994; Hallgren et al. 2001). Rats that were similarly treated 

with pentaBDE for 14 days had increased activities of EROD, MROD, and PROD at ≥18 mg/kg/day 

(lowest tested dose) and increased UDPGT at 36 mg/kg/day (Hallgren et al. 2001). PentaBDE also 

increased hepatic microsomal enzyme activity in maternally-exposed rats and their offspring (Zhou et al. 

2002). Exposure to 1, 10, or 30 mg/kg/day by gavage from gestation day (GD) 6 through postnatal day 

(PND) 21 caused significantly increased hepatic EROD and PROD at ≥10 mg/kg/day in dams (GD 20 and 

PND 22) and offspring (GD 20 and PNDs 4, 14, and 36), as well as increased UDPGT at 30 mg/kg/day in 

dams (GD 20 and PND 22) and offspring (GD 20 and PNDs 4 and 14). 

Renal Effects. 

Polybrominated Biphenyls. No statistically significant correlations were found between serum PBB 

levels and serum levels of blood urea nitrogen (BUN) or creatinine in a study of residents of quarantined 

Michigan farms after the 1973 PBB contamination episode (Kreiss et al. 1982). No other studies were 

located with information pertinent to renal effects in humans after oral exposure to PBBs. 

Studies with animals have shown some renal effects following prolonged exposure to PBBs, but findings 

are generally inconsistent, and the functional significance is uncertain. No exposure-related histological 

changes in kidneys or bladder were observed in rats administered FireMaster FF-1 in a single dose of 

200 or 1,000 mg/kg (Kimbrough et al. 1978b, 1981) and observed for 2–23 months posttreatment 

(Kimbrough et al. 1978a, 1981), or ≤30 mg/kg/day for 2 weeks (Gupta et al. 1981). Gross examination of 

rats exposed to 1,000 mg/kg/day for 2 weeks showed darkened kidneys (Gupta and Moore 1979). Other 

renal information was not reported, but the dosage was lethal. Urinalysis was normal in rats following 

exposure to ≤30 mg/kg/day for 2 weeks (Gupta et al. 1981); urinalysis was not evaluated in the other rat 

studies. Bladder histology, examined in some of the rat studies, was also reported to be normal (Gupta et 

al. 1981; Kimbrough et al. 1981). Kidney histology was not altered in rats exposed to ≤71 mg/kg/day 

octabromobiphenyl mixture for 2 weeks (Lee et al. 1975b; Waritz et al. 1977). Mice exposed to 

≤30 mg/kg/day FireMaster FF-1 for 2 weeks showed no abnormal kidney or bladder histology or 

urinalysis findings (Gupta et al. 1981). Information on acute-duration renal effects in other species was 

not located. 

In intermediate-duration studies with rats, dietary exposure to FireMaster BP-6 for 30 days produced no 

PBB-related alterations in urinalysis indices or BUN at 5 mg/kg/day (highest tested dose) or kidney 
PBBs and PBDEs 181 

5. HEALTH EFFECTS 

histology at ≤10 mg/kg/day (Akoso et al. 1982a; Sleight et al. 1978). However, 5 mg/kg/day FireMaster 

BP-6 in the diet for 3 months caused progressive obsolescence of glomeruli in rats (Bowman's membrane 

was shrunken and glomerular tufts were shrunken, inactive, or had been largely replaced by scar tissue), 

although relative kidney weight, BUN, and renal function tests (clearance of inulin, p-aminohippurate, or 

fractional sodium excretion) were normal (McCormack et al. 1978). Also, in vitro accumulation of 

p-aminohippurate and N-methylnicotinamide, and ammoniagenesis and gluconeogenesis were not 

affected in renal cortical slices from these treated rats. Administration of FireMaster FF-1 by gavage for 

25 weeks caused no renal effects at 0.1 mg/kg/day, but produced chronic progressive nephropathy at 

≥1 mg/kg/day, and more serious histopathology at 10 mg/kg/day (NTP 1983). Renal pathology at the 

10 mg/kg/day dosage included atrophy and edema of glomerular tufts with marked dilation of Bowman's 

capsule and dilation of some renal tubules, with either serous fluid or proteinaceous casts in both cortical 

and medullary regions, and no changes in BUN (NTP 1983). Chronic administration of FireMaster FF-1 

in the diet for up to 104 weeks, however, failed to produce any treatment-related histopathologic changes 

at dosages as high as 1.5 mg/kg/day (NTP 1992). The reason for the inconsistency between this finding 

and the results of the NTP (1983) study is unclear, but the different methods of oral treatment could be a 

factor. 

Intermediate-duration gavage exposure to a higher FireMaster FF-1 dose of 30 mg/kg/day for 4.5 weeks 

caused dilation of Bowman's capsule with serous fluid in rats observed for ≈60 days posttreatment (Gupta 

and Moore 1979); however, rats that were similarly treated (≤30 mg/kg/day for 30 days) but observed 

longer (90 days posttreatment) had normal kidney histology, urinalysis values, and BUN (Gupta et al. 

1981). Rats administered 50 mg/kg/day of an unspecified PBB mixture in the diet for 30 days with no 

posttreatment observation had increased BUN but no changes in urinalysis values or kidney histology 

(Sleight and Sanger 1976). In studies with octabromobiphenyl mixture in rats, dietary exposure to 

≥8 mg/kg/day for 30 days caused hyaline degenerative cytoplasmic changes in kidneys with normal 

urinalysis values (Norris et al. 1975a). This finding is inconsistent with a report of normal kidney 

histology in rats exposed to ≤71 mg/kg/day octabromobiphenyl mixture in the diet for 4 weeks (urinalysis 

not constructed) (Lee et al. 1975b; Waritz et al. 1977); the reason for the discrepancy cannot be discerned 

from the reports. Kidney histology and urinalysis findings were also normal in rats administered 

≤1 mg/kg/day octabromobiphenyl mixture for 8 months (Norris et al. 1975a). In the only study of a 

decabromobiphenyl mixture, urinalysis was normal in rats fed 100 mg/kg/day for 13 weeks (Millischer et 

al. 1980). Comprehensive histology evaluations were performed at this and lower dosages in this study, 

but the liver is the only tissue specifically mentioned as having been examined. Due to the total number 
PBBs and PBDEs 182 

5. HEALTH EFFECTS 

tissues examined (21) and the route of exposure, it is probable that kidneys were examined but were not 

discussed because no histopathologic changes were found. 

In gavage studies with mice, FireMaster FF-1 produced no renal histopathologic changes following 

exposure to ≤10 mg/kg/day for 25 weeks (BUN was normal) (NTP 1983) or ≤30 mg/kg/day for 30 days 

(normal BUN and urinalysis) (Gupta et al. 1981). Dietary exposure to 3.9 mg/kg/day FireMaster FF-1 for 

up to 105 weeks, however, caused an increased incidence of chronic progressive nephropathy in mice; 

this effect was not found at 1.3 mg/kg/day (NTP 1992). Kidney histological alterations were not reported 

in mink exposed to ≤2.4 mg/kg/day FireMaster FF-1 for 313 days (Aulerich and Ringer 1979; Ringer et 

al. 1981). Swine exposed to unspecified PBBs for 16 weeks had increased relative kidney weight at 

≥1 mg/kg/day, but the adversity of this change is unclear since no gross renal pathology was observed 

(≤8 mg/kg/day) and histology was not evaluated (Ku et al. 1978). Monkeys that ingested 18 mg/kg/day 

FireMaster FF-1 for 137 days developed hyperplasia of the bladder epithelium, but histological changes 

in the kidneys were not reported (Allen et al. 1978). Urine alterations in six pregnant cows given 

67 mg/kg/day of FireMaster BP-6 in capsules for up to 60 days included increased protein concentration 

and BUN, and decreased pH and specific gravity (Moorhead et al. 1977). Necropsy following sacrifice 

between days 33 and 66 because of impending death due to poor health showed enlarged, distended, and 

discolored kidneys, extreme dilation of the collecting ducts and convoluted tubules, degenerative changes 

in the tubular epithelium, and congestion with scattered microscopic hemorrhages in the medulla. The 

renal effects in cows appear to be more severe than those generally observed in other species at lethal 

doses. No urinalysis alterations or changes in kidney histology were observed in other cows treated with 

≤0.65 mg/kg/day and examined 1 or 140 days following the end of treatment. 

Polybrominated Diphenyl Ethers. No studies were located regarding renal effects in humans after oral 

exposure to PBDEs. 

Dietary studies of PBDEs in animals observed kidney effects that are mainly attributable to octaBDE. 

Renal effects induced by octaBDE included noninflammatory kidney changes in male rats exposed to 

600 mg/kg/day for 13 weeks (IRDC 1977). The incidence and severity of the kidney lesions (tubule 

regeneration, intratubular casts, and cellular debris occurred in most 600 mg/kg/day males) suggested a 

compound-related effect (IRDC 1977). Another study of octaBDE found no histopathological changes in 

the kidneys of rats exposed to ≤90 mg/kg/day for 90 days (IRDC 1976). Studies of pentaBDE found no 

renal histopathology in rats exposed dietary doses of ≤90 mg/kg/day for 28 days (IRDC 1976) or 

≤100 mg/kg/day for 90 days (WIL Research Laboratories 1984). 
PBBs and PBDEs 183 

5. HEALTH EFFECTS 

Studies of low purity (≈77%) commercial decaBDE mixtures found kidney pathology (hyaline 

degenerative cytoplasmic changes) in male rats exposed to 800 mg/kg/day for 30 days (Norris et al. 1973, 

1975a, 1975b), but not in rats exposed to ≤90 mg/kg/day for 90 days (IRDC 1976) or ≤1.0 mg/kg/day for 

2 years (Kociba et al. 1975; Norris et al. 1975b). Interpretation of this finding is complicated by the fact 

that hyaline degenerative cytoplasmic changes are not uncommon in adult male rats and might be induced 

by a mechanism specific to certain aged male rats. No renal histopathological changes were induced by 

94–97% pure commercial decaBDE in rats exposed to ≤8,000 mg/kg/day for 13 weeks or 

≤2,550 mg/kg/day for 103 weeks (NTP 1986), or in mice exposed to ≤9,500 mg/kg/day for 13 weeks or 

≤7,780 mg/kg/day for 103 weeks (NTP 1986). 

Endocrine Effects. 

Polybrominated Biphenyls. No studies were located regarding endocrine effects in humans after oral 

exposure to PBBs. 

Thyroid effects have been observed in animals treated with PBBs by gavage or diet in acute-, 

intermediate-, and chronic-duration studies. Characteristic changes include decreases in serum levels of 

serum thyroxine (T4) and serum triiodothyronine (T3) hormones, thyroid enlargement, and effects in the 

follicular cells including reduced size, hyperplasia with columnar appearance and papillary projections, 

and accumulation of colloid droplets. In the only acute study that investigated thyroid end points more 

sensitive than histology, rats administered an unspecified PBB mixture for 10 days showed serum T4

levels (T3 not evaluated) that were significantly reduced (p≤0.05) at ≥3 mg/kg/day, but not at 1 mg/kg/day 

(Allen-Rowlands et al. 1981). The reduction in T4 levels was both dose- and time-dependent as shown by 

20-day results discussed below with intermediate-duration studies. Based on the NOAEL for decreased 

serum T4, an acute oral MRL of 0.01 mg/kg/day was calculated as indicated in the footnote to Table 5-2 

and discussed in Chapter 3 and Appendix A. A single ≤286 mg/kg dose of an unspecified PBB mixture 

caused no change in 4-hour thyroidal 131I uptake and incorporation into thyroglobulin in rats (AllenRowlands et al. 1981). No thyroid histological alterations were observed in rats in acute-duration studies 

with FireMaster FF-1, even with a single dose ≤1,000 mg/kg and up to 2 years posttreatment observation 

(Kimbrough et al. 1978b, 1981) or dosages of ≤1,000 mg/kg/day for 2 weeks (Gupta and Moore 1979; 

Gupta et al. 1981). The only information on thyroid effects of acute exposure to octabromobiphenyl 

mixture is a lack of histological changes in rats administered ≤71 mg/kg/day for 2 weeks (Lee et al. 

1975b; Waritz et al. 1977). The only information on acute-duration thyroid effects of PBBs in species 
PBBs and PBDEs 184 

5. HEALTH EFFECTS 

other than rat is the normal histologic integrity of the thyroid in mice at FireMaster FF-1 dosages of 

≤30 mg/kg/day for 2 weeks (Gupta et al. 1981). 

In intermediate-duration studies with rats, serum levels of T3 or T4 decreased at FireMaster dosages as 

low as 0.3 mg/kg/day FireMaster FF-1 for 25 weeks (NTP 1983), 0.45 mg/kg/day FireMaster BP-6 for 

7 months (Byrne et al. 1987), 5 mg/kg/day FireMaster BP-6 for 30 days (Akoso et al. 1982b) or 

1 mg/kg/day of an unspecified PBB mixture for 20 days (Allen-Rowlands et al. 1981). In the latter study, 

8–11 rats were evaluated after exposure to 1, 3, or 6 mg/kg/day for 20 days. Other thyroid effects in these 

rats included significantly increased absolute thyroid weight at ≥3 mg/kg/day (not evaluated at 

1 mg/kg/day), and increased plasma TSH levels, increased 5-hour thyroid uptake of 131I and deceased 

incorporation of 131I into monoiodotyrosine (MIT) at 6 mg/kg/day (Allen-Rowlands et al. 1981). No 

effects on incorporation of 131I into diiodotyrosine (DIT), T3, or T4 were observed. Serum T4 levels were 

also reduced at ≥1 mg/kg/day in rats exposed for 20 days and evaluated after being placed on restricted 

food intake for ≥2 months following treatment (Allen-Rowlands et al. 1981). Rats administered 

2.5 mg/kg/day of an unspecified hexabromobiphenyl mixture for 7 months showed no significant changes 

in serum T3, but serum T4 was not evaluated (Sepkovic and Byrne 1984). Thyroid ultrastructural changes 

were produced in rats by FireMaster BP-6 dosages as low as 0.05 mg/kg/day for 30–35 days (Akoso et al. 

1982b; Kasza et al. 1978a), and histologic changes of the thyroid were observed at ≥5 mg/kg/day 

FireMaster BP-6 for 30 days (Sleight et al. 1978) and ≥30 mg/kg/day FireMaster FF-1 for 4.5 weeks 

(Gupta and Moore 1979). In the study that evaluated thyroid effects at the lowest dose, rats were 

administered estimated doses of 0.05, 0.5, or 5 mg/kg/day FireMaster BP-6 in the diet for 30 days (Akoso 

et al. 1982b). Effects were dose-dependent and included increased number and decreased size of follicles 

(especially at the peripheral location) at ≥0.05 mg/kg/day, follicles with epithelial tall columnar 

appearance and some papillary projections in the lumen at ≥0.5 mg/kg/day, and extensive follicular 

changes (hyperplasia and hypertrophy of follicular cells, prominent, and numerous papillary projections), 

increased relative thyroid weight, and decreased serum T3 and T4 at 5 mg/kg/day. Chronic exposure to 

≤1.5 mg/kg/day FireMaster FF-1 for up to 104 weeks caused no thyroid histological alterations, but 

ultrastructure and serum thyroid hormones were not assayed (NTP 1992). 

In the only intermediate-duration rat study of octabromobiphenyl mixture that assessed thyroid hormones, 

a dose of 2.5 mg/kg/day for 7 months produced no significant changes in serum T3, but serum T4 was not 

evaluated (Sepkovic and Byrne 1984). The histologic integrity of the thyroid was normal in rats fed 

octabromobiphenyl mixture at dosages as high as 1 mg/kg/day for 8 months (Norris et al. 1975a), 

2.5 mg/kg/day for 7 months (Sepkovic and Byrne 1984), and 71 mg/kg/day for 4 weeks (Lee et al. 1975b; 
PBBs and PBDEs 185 

5. HEALTH EFFECTS 

Waritz et al. 1977), although ≥8 mg/kg/day for 30 days induced dose-related thyroid hyperplasia (Norris 

et al. 1975a). An explanation for the discrepancy in the octabromobiphenyl mixture NOAELs of 

≤71 mg/kg/day and LOAELs of ≥8 mg/kg/day is not apparent, particularly since treatment durations were 

similar, methods of treatment (diet) and animal strain and sex (male) were the same, and only the NOAEL 

study appears to have observed the animals (for 2–18 weeks) posttreatment. 

Effects on the adrenal gland also have been observed in animals exposed to PBBs. As found for thyroid 

as discussed above, acute-duration exposure to FireMaster FF-1 produced no changes in rat adrenal 

histology following a single dose as high as 1,000 mg/kg (Gupta et al. 1981; Kimbrough et al. 1978b, 

1981). Dosages of 1,000 mg/kg/day FireMaster FF-1 for 2 weeks caused gross adrenal damage (darkened 

glands) in rats, but ≤30 mg/kg/day caused no gross or histologic damage (Gupta and Moore 1979; Gupta 

et al. 1981). The only information on acute-duration adrenal effects of PBBs in species other than rat is 

normal adrenal histology in mice at FireMaster FF-1 dosages of ≤30 mg/kg/day for 2 weeks (Gupta et al. 

1981). No acute-duration studies of PBBs measure serum corticosteroid levels. In intermediate-duration 

studies, serum corticosterone B, dehydroepiandrosterone (DHEA), and dehydroepiandrosterone sulfate 

(DHS) decreased in rats fed ≥0.25 mg/kg/day FireMaster BP-6 for 5–7 months, but not 0.05 mg/kg/day 

(Byrne et al. 1988). Serum corticosterone levels and adrenal weight did not change in rats exposed to 

≤6 mg/kg/day of an unspecified PBB mixture for a shorter duration of 20 days (other adrenal hormones 

were not evaluated) (Castracane et al. 1982). Adrenal histology was not evaluated in these studies, but no 

treatment-related alterations were observed in rats in other intermediate-duration studies with FireMaster 

BP-6 or FireMaster FF-1 (Akoso et al. 1982b; NTP 1983; Sleight and Sanger 1976; Sleight et al. 1978), 

except at lethal dosages (Gupta and Moore 1979), or in a chronic study with FireMaster FF-1 (NTP 

1992). Necropsies of rats treated with 100–1,000 mg/kg/day FireMaster FF-1 for 4.5 weeks showed 

darkened adrenals (Gupta and Moore 1979). In the only rat study of octabromobiphenyl mixture that 

examined the adrenal gland, 2.5 mg/kg/day for 7 months produced no changes in relative adrenal weight; 

histology or serum corticosteroids were not evaluated (Sepkovic and Byrne 1984). Intermediate- or 

chronic-duration studies with FireMaster FF-1 in mice showed no adrenal histological effects at 

≤3.9 mg/kg/day for up to 105 weeks (NTP 1992), ≤10 mg/kg/day for 25 weeks (NTP 1983), or 

≤30 mg/kg/day for 30 days (Gupta et al. 1981). 

Polybrominated Diphenyl Ethers. Plasma levels of the congener 2,2',4,4'-tetraBDE (BDE 47) and 

various other persistent organohalogen compounds (non-PBDEs), as well as hormone levels (free and 

total T3 and T4, thyroid stimulating hormone [TSH], free testosterone, follicle-stimulating hormone, 

lutenizing hormone, and prolactin), were analyzed in 110 men who consumed varying amounts of fatty 
PBBs and PBDEs 186 

5. HEALTH EFFECTS 

fish (0–32 meals per month) from the Baltic Sea (Hagmar et al. 2001). There was a weak negative 

correlation between BDE 47 and plasma TSH after age adjustment, but the congener could not explain 

more than 10% of the variance in TSH (r2

=0.10, p<0.001). The fact that BDE 47 could only explain 10% 

of the variance in TSH is not surprising due to the occurrence of PCBs and other likely similarly acting 

compounds in the Baltic fisherman. 

Hyperplasia of the thyroid was observed in rats and mice following repeated dietary exposures to 

decaBDE. Thyroid follicular cell hyperplasia was increased in male B6C3F1 mice that were exposed to 

≥94% pure commercial decaBDE for 103 weeks (NTP 1986). Incidences of the lesion were 2/50 (4%), 

10/50 (20%), and 19/50 (38%) in the 0, 3,200, and 6,650 mg/kg/day dose groups of this study. Slight 

increases in follicular cell tumors that were considered to be equivocal evidence of thyroid 

carcinogenicity were also observed in the male mice (see Section 5.2.2.7, Cancer). No decaBDE-related 

histopathological changes in the thyroid were found after 103 weeks of exposure to ≤7,780 mg/kg/day in 

female mice, ≤2,240 mg/kg/day in male Sprague-Dawley rats, or ≤2,550 mg/kg/day in female rats (NTP 

1986). Dose-related increases in thyroid hyperplasia were reported for male Sprague-Dawley rats 

exposed to 80 and 800 mg/kg/day for 30 days (Norris et al. 1973, 1975b), although not in rats exposed to 

≤90 mg/kg/day for 90 days, rats exposed to ≤8,000 mg/kg/day for 13 weeks, or mice exposed to 

≤9,500 mg/kg/day for 13 weeks (IRDC 1976; NTP 1986). The occurrence of thyroid hyperplasia in the 

rats exposed to ≥80 mg/kg/day for 30 days could be related to the low purity composition of the older 

commercial decaBDE mixture tested by Norris et al. (1973, 1975a, 1975b) (i.e., 77.4% decaBDE, 21.8% 

nonaBDE, and 0.8% octaBDE, compared to the ≥94% decaBDE composition used in the NTP studies). 

Thyroid function was not tested in any of the intermediate- and chronic-duration studies of decaBDE, 

although an acute study in rats (Zhou et al. 2001), summarized in the following paragraph, found that a 

>98% pure decaBDE product (DE-83R) caused no serum changes in thyroid hormones. Studies of 

commercial octa- and pentaBDE mixtures and individual constituent congeners of these mixtures, also 

summarized below, indicate that the rat and mouse thyroid is particularly sensitive to the lower 

brominated BDEs. The human relevance of the animal thyroid data is unclear due to evidence indicating 

that the main thyroid effect, decreased serum T4, occurs by a mechanism that is not clearly relevant to 

humans (see Section 5.5.2 Mechanisms of Toxicity). 

Thyroid hormone levels were determined in weanling (28-day-old) female Long-Evans that were treated 

by gavage for 4 days with commercial mixtures of decaBDE (DE-83R) or octaBDE (DE-79) in doses of 

0.3, 1, 3, 10, 30, 60, or 100 mg/kg/day, or pentaBDE (DE-71) in doses of 0.3, 1, 3, 10, 30, 100, or 
PBBs and PBDEs 187 

5. HEALTH EFFECTS 

300 mg/kg/day (Zhou et al. 2001). The animals were sacrificed on the day after the last exposure and 

evaluated for changes in serum levels of total T4, total T3, and TSH. DecaBDE caused no changes in 

levels of any of the thyroid hormones. OctaBDE induced a dose-related reduction in serum T4 levels with 

statistically significant (p<0.05) decreases occurring at ≥10 mg/kg/day and a 70% maximum decrease 

compared to controls at the highest dose of 100 mg/kg/day. Serum total T3 levels were significantly 

reduced at ≥60 mg/kg/day with a maximum reduction of 25% at 100 mg/kg/day. PentaBDE caused doserelated reductions in serum T4 levels with significant decreases occurring at ≥30 mg/kg/day and an 80% 

maximum decrease compared to controls at the highest dose of 300 mg/kg/day. Serum total T3 levels 

were significantly reduced at ≥100 mg/kg/day with a maximum reduction of 30% at 300 mg/kg/day. 

Neither octaBDE nor pentaBDE caused exposure-related changes in serum TSH concentrations. 

Benchmark dose (BMD) analysis of the octaBDE data found that the BMD and BMDL (95% lower 

confidence limit on the BMD) resulting in a 20% reduction in thyroid hormones (LED20) were 9.25 and 

5.29 mg/kg/day, respectively, for serum T4 and 53.38 and 11.98 mg/kg/day, respectively, for serum T3. 

For pentaBDE, the respective BMD and BMDL resulting in 20% reduced hormone levels were 12.74 and 

6.95 mg/kg/day for serum T4 and 32.94 and 8.56 mg/kg/day for serum T3. 

Serum levels of thyroid hormones were not evaluated in any other study of octaBDE. Thyroid 

hyperplasia was equivocally increased in male CD rats that were exposed to 90 mg/kg/day of octaBDE 

(unspecified mixture) in the diet for 28 days (IRDC 1976); incidences of slight or moderate hyperplasia 

were 0/5, 0/5, and 3/5 at 0, 9, and 90 mg/kg/day, respectively. CD rats that were exposed to octaBDE in 

estimated dietary doses of 5, 50, or 600 mg/kg/day (males) or 7, 70, or 750 mg/kg/day (females) for 

13 weeks had increased absolute and relative thyroid weights at ≥50/70 mg/kg/day (IRDC 1977). The 

thyroid weight increases were still observed at 8 weeks postexposure in the 600/750 mg/kg/day groups 

and were concluded to be likely compound-related. Most of the follicles in the thyroids of 4/35 males at 

600 mg/kg/day and 1/35 females at 750 mg/kg/day had epithelium that was tall columnar rather than the 

normal cuboidal type. This effect was considered to be a very slight but probably compound-related 

histological change. The thyroid glands were considered within the range of normal morphology at 

8 weeks postexposure in the 600/750 mg/kg/day rats as well as in the lower dose groups (IRDC 1977). 

The 4-day study in weanling rats summarized above (Zhou et al. 2001) is not the only study of the 

commercial pentaBDE mixture DE-71 that found reduced thyroid hormone levels in maternally-exposed 

rats and their offspring (Zhou et al. 2002). Long-Evans rats were administered 1, 10, or 30 mg/kg/day of 

pentaBDE (DE-71) in corn oil by gavage from GD 6 through PND 21. Total concentrations of serum T4

and T3 were evaluated in dams on GD 20 and PND 22 and offspring were evaluated on GD 20 and PNDs 
PBBs and PBDEs 188 

5. HEALTH EFFECTS 

4, 14, 36, and 90. Serum T4 was significantly (p<0.05) reduced compared to controls in dams at 

30 mg/kg/day on GD 20 and PND 22 (48 and 44% reduced, respectively), and in offspring at 

≥10 mg/kg/day on GD 20 (≥15% reduced) and PNDs 4 and 14 (50 and 64% reduced at 10 and 

30 mg/kg/day, respectively). The effect on serum T4 concentrations in the offspring returned to control 

levels by PND 36. The NOAEL for the effect on serum T4 levels (1 mg/kg/day) was used as the basis for 

the acute-duration MRL for oral exposure to tetra- to heptaBDEs as indicated in the footnote to Table 5-3 

and discussed in Chapter 4 and Appendix A. The BMD and BMDL resulting in a 20% reduction in 

serum T4 levels were reported to be 2.36 and 0.94 mg/kg/day, respectively. There were no exposurerelated changes in serum T3 levels in the dams or offspring, or any significant effects of treatment on litter 

size, sex ratio, or nonneurodevelopmental measures of offspring viability and growth as discussed in 

Section 5.2.2.6 (Developmental Effects). 

Reduced serum T4 levels additionally occurred in adult offspring of rats that were exposed to commercial 

pentaBDE mixture DE-71 during gestation and lactation in two studies incompletely reported as abstracts 

(Taylor et al. 2002, 2003). In one study, rats were orally dosed with 0, 1, 10, or 30 mg/kg/day of DE-71 

in corn oil from GD 6 through PND 21 (Taylor et al. 2002). Evaluation of the offspring at various ages 

(not specified) showed effects that included reduced serum T4 levels; specific data were not reported, 

although the BMD and BMDL for reduced T4 were 2.3 and 0.9 mg/kg/day, respectively. In the other 

study, rats were similarly administered 1-100 mg/kg/day of DE-71 on GD 6 to PND 21 (Taylor et al. 

2003). Measurement of serum levels of thyroid hormones in offspring showed dose-related decreases in 

T4 at 5, 30, and 100 mg/kg/day on PND 5 (73.3, 49.3, and 43.5% of controls) and PND 14 (75.6, 33.6, 

and 29.9% of controls). The serum T4 decreases in both studies were accompanied by increases in 

hepatic metabolism (EROD, PROD, UDPGT, and glucuronidation). Other effects included neurobehavioral and reproductive developmental changes as summarized in Sections 5.2.2.4 (Neurological 

Effects) and 5.2.2.5 (Reproductive Effects). 

Commercial pentaBDE mixture DE-71 was tested under EPA’s Endocrine Disruptor Screening Program 

(EDSP) using male and female rat pubertal protocols for detecting thyroid active agents (Laws et al. 

2003; Stoker et al. 2003). These studies are incompletely reported as abstracts. End points in both 

studies included pubertal development (time to preputial separation in males and vaginal opening in 

females), body and reproductive tissue weights, thyroid function (serum levels of T4, T3, and TSH), and 

liver microsomal enzyme induction (EROD, PROD, and UDPGT activity). Both sexes showed delays in 

reproductive development that were accompanied by changes in thyroid hormone levels and hepatic 

metabolism. In the study with males, Wistar rats were gavaged with 0, 3, 30, or 60 mg/kg/day doses in 
PBBs and PBDEs 189 

5. HEALTH EFFECTS 

corn oil for 5 days on PNDs 24–28, or for 31 days on PNDs 23–53 (Stoker et al. 2003). Pubertal 

development was delayed as shown by significant increases in the age of preputial separation at 30 and 

60 mg/kg/day (2.0- and 2.3-day delay, respectively). Seminal vesical and ventral prostate weights were 

reduced at 60 mg/kg/day with no significant changes in testis or epididymal weights or serum testosterone 

levels. Thyroid hormone measurements showed reductions in serum T4 at 30 and 60 mg/kg/day (74 and 

81% lower than controls) after 5 days of exposure, serum T4 at 3, 30, and 60 mg/kg/day (20, 80, and 86% 

lower) after 31 days, and serum T3 at 30 and 60 mg/kg/day (25 and 20% lower) after 31 days, and 

increased serum TSH at 30 and 60 mg/kg/day (64 and 113% higher than controls) after 31 days. Hepatic 

effects included increased relative liver weight and microsomal EROD, PROD, and UDPGT activity at 

30 and 60 mg/kg/day at both time points. In the study with females, Wistar rats were gavaged with DE71 in corn oil in doses of 0, 3, 30, or 60 mg/kg/day for 5 days on PNDs 22–26, or for 20 days on 

PNDs 22–41 (Laws et al. 2003). The 60 mg/kg/day dose caused a small but statistically significant delay 

in the age of vaginal opening (1.8-day delay compared to controls). Other effects included a significant 

decrease in serum T4 at ≥30 mg/kg/day at both time points, a linear trend for increased TSH (1.6-fold 

increase at 60 mg/kg/day) after 20 days of exposure, and increased relative liver weight and microsomal 

EROD, PROD, and UDPGT activity at ≥30 mg/kg/day after at both time points. The results of both 

studies indicate that the changes in thyroid hormone levels were related to induction of liver metabolic 

enzymes. 

End points assessed in a comprehensive 90-day feeding study of a commercial pentaBDE (DE-71) 

mixture in male and female Sprague-Dawley rats included serum T3 and T4 levels (TSH not measured) 

and thyroid histology (WIL Research Laboratories 1984). Effects observed in both sexes included 

significantly reduced plasma T4 levels at ≥10 mg/kg/day and increased follicular cell hyperplasia at 

100 mg/kg/day. Incidences of follicular cell hyperplasia in the 0, 2, 10, and 100 mg/kg/day dose groups 

of this study were 0/10, 2/10, 2/10, and 5/10 in males and 0/10, 0/10, 1/10, and 4/10 in females. The 

thyroid hyperplasia was mild and transient as it was characterized as very slight in severity at all doses 

and was no longer observed at 24 weeks postexposure in any animals. Thyroid hyperplasia was 

equivocally increased in male CD rats that were exposed to 90 mg/kg/day of pentaBDE (unspecified 

mixture) in the diet for 28 days (IRDC 1976). Incidences of slight or moderate hyperplasia in the 0, 9, or 

90 mg/kg/day dose groups of this study were 0/5, 1/5, and 3/5 males, respectively; no increases were seen 

in females. 

Serum total T4 concentrations were significantly reduced in female C57BL/6J mice that were given a 

single gavage dose of commercial pentaBDE mixture DE-71 as low as 0.8 mg/kg (lowest tested dose) 
PBBs and PBDEs 190 

5. HEALTH EFFECTS 

(Fowles et al. 1994). Reductions in serum T4 were found in four of five dose groups (i.e., 0.8, 4, 20, and 

500 mg/kg, but not 100 mg/kg), but did not occur in a dose-dependent manner. This lack of dosedependency might be explained by the number of animals per group, which was relatively small (n=6). 

Another single-dose gavage study, this time in rats (Hakk et al. 2002), reported transient increases in total 

T4 plasma levels, but the small (n=3) number of animals per group and the lack of reported tests of 

statistical significance make the interpretation of the findings unclear. Evaluation of mice exposed to 18, 

36, or 72 mg/kg/day of pentaBDE by daily gavage for 14 days showed significantly reduced total and free 

T4 levels at ≥18 mg/kg/day (Fowles et al. 1994). Serum T3 and TSH levels and thyroid histology were not 

evaluated in these single-dose and 14-day mouse studies. Serum levels of total and free T4 were also 

reduced in female Sprague-Dawley rats and female B6C3F1 mice that were exposed to ≥18 mg/kg/day of 

the commercial pentaBDE mixtures Bromkal 70 or Bromkal 70-5 DE, or the congener 2,2',4,4'-tetraBDE 

(BDE 47), by daily gavage for 14 days (Darnerud and Sinjari 1996; Hallgren and Darnerud 1998; 

Hallgren et al. 1998, 2001). Exposure to the mixtures or congener did not cause decreases in serum T4 at 

doses ≤6 mg/kg/day, or any changes in serum TSH or thyroid histology at any tested level 

(≤36 mg/kg/day) in either species. The decreases in serum T4 were associated with reduced ex vivo

binding of T4 to the plasma thyroid hormone transporter protein transthyretin, as well as with induction of 

hepatic microsomal enzymes (EROD, MROD, PROD, and UDPGT) (Hallgren and Darnerud 1998). A 

limited amount of information is available on hormonal effects of PBDEs other than thyroid. There were 

no clear chemical-related changes in serum corticosterone levels in female mice that were exposed to a 

commercial pentaBDE mixture (DE-71) in doses of 18, 36, or 72 mg/kg/day by daily gavage for 14 days 

(Fowles et al. 1994). As reported in a study abstract (Kuriyama and Chahoud 2003), serum levels of 

testosterone and LH were unchanged in adult male offspring of Wistar rats that were maternally exposed 

to a single 60 or 300 μg/kg oral dose of the penta congener BDE 99 on day 6 of gestation. Although there 

were no effects on these male reproductive hormones, spermatid and sperm numbers were reduced as 

summarized in Section 5.2.2.6 Developmental Effects. 

Dermal Effects. 

Polybrominated Biphenyls. Limited human data from an epidemiological study provide suggestive 

evidence that oral exposure to PBBs may produce skin disorders in humans, but do not provide 

information regarding dose-response relationships. Symptoms of skin disorders (rashes, acne, increased 

sensitivity to the sun, darkening or thickening of the skin, discoloration or deformity of fingernails or 

toenails, slow healing of cuts) were reported with greater frequency in a group of 406 Michigan residents 

probably exposed to PBBs than in a group of 153 likely unexposed residents, but no association was 

evident between serum PBB levels and prevalence of skin disorders (Anderson et al. 1978c). In a medical 
PBBs and PBDEs 191 

5. HEALTH EFFECTS 

history survey study conducted in 1976, symptoms of skin disorders (peeling and scaling, erythema, hair 

loss, increased nail growth, increased sweating) experienced during the previous 3 years were reported at 

higher prevalence rates in a group of 321 Michigan residents from quarantined farms and in a group of 

177 nonquarantined farm residents than in a group of 149 nonexposed Wisconsin residents (Chanda et al. 

1982). Physical examination of the combined group of Michigan residents revealed alopecia in 4% of the 

subjects compared to no occurrence of alopecia in the control group. 

In animals, repeated exposures to PBBs produced characteristic dermal changes in certain species, 

particularly monkeys, but generally not in haired rodents. No exposure-related histological changes were 

observed in the skin, salivary glands, or eyes of rats administered a single dose of 200 mg/kg FireMaster 

FF-1 and observed for 18–22 months (Kimbrough et al. 1981). Rats and mice exposed to ≤30 mg/kg/day 

FireMaster FF-1 for 2 weeks showed no histological alterations in pinnae, ear canals, or salivary glands, 

but examination of skin was not performed (Gupta et al. 1981). In intermediate- and chronic-duration 

studies, rats and/or mice exposed to FireMaster FF-1 or FireMaster BP-6 dosages of ≤30 mg/kg/day by 

gavage for 30 days (Gupta et al. 1981), ≤10 mg/kg/day in the diet for 30 days (Akoso et al. 1982a), 

≤10 mg/kg/day by gavage for 25 weeks (NTP 1983), or ≤3.9 mg/kg/day in the diet for up to 105 weeks 

(Chhabra et al. 1993; NTP 1992) showed no histopathological changes in skin, pinnae, ear canals, or 

salivary glands. Xerophthalmia (extreme dryness of the conjunctiva, with keratinization of epithelium 

following chronic conjunctivitis) was observed in rats after 82 days of dietary exposure to 5 mg/kg/day 

FireMaster BP-6 (Darjono et al. 1983). Alopecia, loss of eyelashes, generalized subcutaneous edema, dry 

scaly skin, and periorbital edema developed in monkeys exposed to FireMaster FF-1 in dosages of 

≥0.73 mg/kg/day for ≥25 weeks (two animals) or 18 mg/kg/day for 137 days (one animal) (Allen et al. 

1978; Lambrecht et al. 1978). Histological examination, performed only in the monkey exposed to 

18 mg/kg/day, showed atrophy and squamous metaplasia of sebaceous glands and keratinization of hair 

follicles (Allen et al. 1978). Dermatosis on the ventral surface was a clinical sign in two of four swine 

administered 8 mg/kg/day FireMaster FF-1 for 16 weeks (Ku et al. 1978). No additional information was 

reported on the dermatosis (a nonspecific term used to denote any cutaneous lesion or group of lesions), 

and histologic examinations were not completed. 

Polybrominated Diphenyl Ethers. No studies were located regarding dermal effects in humans after oral 

exposure to PBDEs. 
PBBs and PBDEs 192 

5. HEALTH EFFECTS 

Histopathological examinations showed no dermal changes in rats following dietary exposure to 

≤750 mg/kg/day of octaBDE for 13 weeks (IRDC 1977) or ≤100 mg/kg/day of pentaBDE for 90 days 

(WIL Research Laboratories 1984). 

Ocular Effects. 

Polybrominated Biphenyls. No studies were located regarding ocular effects in humans after oral 

exposure to PBBs. 

Occasional eye discharge was observed in pregnant cows that were administered 67 mg/kg/day of 

FireMaster BP-6 in capsules for up to 60 days (Moorhead et al. 1977). Necropsy following sacrifice 

between days 33 and 66 because of impending death due to poor health showed hyperkeratosis of the 

eyelids and squamous metaplasia with keratin cysts in the tarsal glands in five of six animals. No ocular 

effects were observed in other cows treated with ≤0.65 mg/kg/day and examined 1 or 140 days following 

the end of treatment. Histological changes were not observed in the eyes of rats exposed to FireMaster 

FF-1 for 2 weeks (Gupta et al. 1981), or in rats and mice treated by gavage (NTP 1983) or fed FireMaster 

FF-1 for up to 105 weeks (Chhabra et al. 1993; NTP 1992). 

Polybrominated Diphenyl Ethers. No studies were located regarding ocular effects in humans after oral 

exposure to PBDEs. 

Histopathological examinations showed no ocular effects in rats following dietary exposure to 

≤1.0 mg/kg/day of 77.4% decaBDE (containing 21.8% nonaBDE and 0.8% octaDBE) for 2 years (Kociba 

et al. 1975; Norris et al. 1975b), ≤750 mg/kg/day of octaBDE for 13 weeks (IRDC 1977), or 

≤100 mg/kg/day of pentaBDE for 90 days (WIL Research Laboratories 1984). 

Body Weight Effects. 

Polybrominated Biphenyls. No studies were located regarding body weight effects in humans after oral 

exposure to PBBs. 

Reduced body weight was observed in various species following acute oral administration of relatively 

high doses of PBBs; this effect is most evident with repeated exposure. In general, decreases in food and 

water intake are not sufficient to account for decreases in body weight. Effects on body weight can be 

quite pronounced following intermediate- and chronic-duration exposure, constituting a wasting 
PBBs and PBDEs 193 

5. HEALTH EFFECTS 

syndrome manifested by weight loss and depletion of body fat. In acute-duration studies with rats, a 

single 1,000 mg/kg dose of FireMaster FF-1 caused decreased weight gain during the following 2 years 

(Kimbrough et al. 1981), and a single 800 mg/kg dose of FireMaster BP-6 during pregnancy caused 

maternal weight loss (Beaudoin 1977). Single FireMaster doses of 400 (BP-6) or 500 (FF-1) mg/kg/day 

did not affect body weight in rats (Beaudoin 1977; Kimbrough et al. 1980). Administration of 

1,000 mg/kg/day FireMaster FF-1 or 130 mg/kg/day FireMaster BP-6 for 2 weeks produced decreased 

weight gain or weight loss in rats and mice, respectively (Corbett et al. 1978; Fraker 1980; Fraker and 

Aust 1978; Gupta and Moore 1979), but 5 mg/kg/day FireMaster BP-6 for 10 days had no effect on body 

weight in rats (Raber and Carter 1986). A single ≤2,000 mg/kg dose or two daily 3,000 mg/kg doses of 

octabromobiphenyl mixture had no effect on body weight gain in rats observed for the following 14–

28 days (Lee et al. 1975a; Norris et al. 1975a). No changes in body weight were produced in rats exposed 

to ≤71 mg/kg/day octabromobiphenyl mixture for 2 weeks (Lee et al. 1975b; Waritz et al. 1977). In 

intermediate-duration studies, decreased body weight gain and/or weight loss has been observed in rats at 

dosages as low as 3 mg/kg/day FireMaster FF-1 for 6 months (Luster et al. 1980; NTP 1983), 

5 mg/kg/day FireMaster BP-6 for 1–3 months (McCormack et al. 1978; Sleight et al. 1978), and 30–

50 mg/kg/day FireMaster FF-1 for 4.5–5 weeks (Gupta and Moore 1979; Sleight and Sanger 1976). 

FireMaster FF-1 dosages ≥100 mg/kg/day for 4.5 weeks (lethal doses) caused weight loss and emaciation 

in rats (Gupta and Moore 1979). Final body weights were decreased ≥11–28% in rats exposed to 0.5 or 

1.5 mg/kg/day FireMaster FF-1 for up to 104 weeks (NTP 1992). No body weight changes were 

observed in rats fed a decabromobiphenyl mixture at dosages as high as 100 mg/kg/day for 13 weeks 

(Millischer et al. 1980). 

In mice exposed to FireMaster FF-1, estimated dosages of 10 mg/kg/day for 25 weeks (NTP 1992) and 

21.7 mg/kg/day for 6 weeks (Loose et al. 1981) decreased the rate of weight gain. Chronic exposure to 

≤3.9 mg/kg/day FireMaster FF-1 for up to 105 weeks, however, did not produce adverse effects on mouse 

body weight (NTP 1992). Guinea pig, mink, and monkey seem to be particularly sensitive species, as 

indicated by pronounced weight loss in guinea pigs from ingestion of 4 mg/kg/day of unspecified PBBs 

for 30 days (Sleight and Sanger 1976), decreased weight gain in mink at FireMaster FF-1 dosages as low 

as 0.39 mg/kg/day with weight loss at 1.86 mg/kg/day (Aulerich and Ringer 1979; Ringer et al. 1981), 

and weight loss in monkeys at FireMaster FF-1 dosages as low as 0.73 mg/kg/day for 25–50 weeks (Allen 

et al. 1978; Lambrecht et al. 1978). Monkeys that ingested an estimated FireMaster FF-1 dosage of 

0.012 mg/kg/day for 66 weeks lost weight (Lambrecht et al. 1978). Food intake and body weight gain 

were reduced in pregnant cows after 4 and 20 days administration of 67 mg/kg/day FireMaster BP-6 in 

capsules (Moorhead et al. 1977). This dosage was lethal because death was impending between days 33 
PBBs and PBDEs 194 

5. HEALTH EFFECTS 

and 66 (treatment duration was 60 days). There were no effects on food intake or body weight in cows 

treated with ≤0.65 mg/kg/day and observed 1 or 140 days following the end of treatment. 

Polybrominated Diphenyl Ethers. No studies were located regarding body weight effects in humans 

after oral exposure to PBDEs. 

DecaBDE had no effect on body weight gain in rats and mice that were exposed to dietary doses of 

≤16,000 and ≤19,000 mg/kg/day, respectively for 14 days, ≤8,000 and ≤9,500 mg/kg/day, respectively, 

for 13 weeks (NTP 1986), or ≤2,550 and ≤7,780 mg/kg/day, respectively, for 103 weeks (NTP 1986). 

Dietary ingestion of 77.4% decaBDE mixture (containing 21.8% nonaBDE and 0.8% octaDBE) similarly 

caused no body weight changes in rats exposed to ≤800 mg/kg/day for 30 days or ≤1.0 mg/kg/day for 

2 years (Kociba et al. 1975; Norris et al. 1973, 1975a, 1975b). 

OctaBDE did not affect body weight in rats exposed to dietary doses of ≤90 mg/kg/day for 28 days 

(IRDC 1976), although exposure to ≥600 mg/kg/day for 13 weeks caused ≥12% decreases in weight gain 

(IRDC 1977). There were no pentaBDE-related body weight changes in rats exposed to ≤90 mg/kg/day 

for 28 days (IRDC 1976) or mice exposed to ≤72 mg/kg/day pentaBDE for 14 days (Fowles et al. 1994), 

although the rate of weight gain was reduced in rats exposed to ≤100 mg/kg/day pentaBDE for 90 days 

(WIL Research Laboratories 1984). 

5.2.2.3 Immunological and Lymphoreticular Effects 

Polybrominated Biphenyls. Numerous reports have been published regarding the immunological 

competence of individuals exposed to PBBs in the Michigan feed contamination episode. Due to the 

relatively high number of published reports and to the fact that often different groups of investigators 

appear to have examined the same cohort, only representative studies are discussed below. 

Immunological parameters were compared between a group of 45 adult Michigan dairy farmers and their 

families who were exposed for periods ranging from 3 months to 4 years and two groups of control 

individuals not known to have been exposed to PBBs (Bekesi et al. 1978, 1979). In 27 of the 

45 Michigan subjects, the peripheral blood lymphocytes responded within a normal range to 

phytohemagglutinin (PHA) and PWM mitogen-induced lymphoblastogenesis, but had reduced reactivity 

in mixed leukocyte cultures relative to controls. In the remaining 18 Michigan subjects, the response to 

PHA and PWM and the reactivity to mixed leukocyte cultures was significantly reduced (p<0.00001) 
PBBs and PBDEs 195 

5. HEALTH EFFECTS 

relative to controls. Assays for membrane markers of peripheral blood lymphocytes showed significant 

reductions in markers in the Michigan subpopulation with abnormal lymphoblastogenesis. Both 

Michigan subpopulations had a significant increase in the number of lymphocytes without detectable 

surface markers, relative to controls. The number of markers for monocytes was not significantly 

different among the groups studied. There were no significant differences in serum PBB levels between 

the two Michigan subsets. No consistent correlation could be demonstrated between lymphocyte function 

and PBB plasma concentration. 

Reexamination of a group of 40 Michigan farmers 5 years after the first examination (Bekesi et al. 1985; 

Roboz et al. 1985) showed that the number of T-lymphocytes and the lymphocyte response to stimulation 

with PHA were altered to the same extent reported 5 years earlier (Bekesi et al. 1978). 

In a similar study, Michigan subjects were classified into three groups according to their serum PBB 

levels: high (>300 ppb), low (<1–11 ppb), and unexposed (controls) (Silva et al. 1979). The percentage 

of subjects that complained of recurrent infection was similar in the two exposed groups (about 20%). 

The total leukocyte count, percentage of lymphocytes, and percentage of subpopulations of T- and Blymphocytes were similar among the three groups. Mean spontaneous lymphocyte transformation and 

lymphocyte mitogenic responsiveness to stimulation with three different mitogens were not significantly 

different among the three groups. Furthermore, there was no correlation between a poor mitogenic 

response and low numbers of T-lymphocytes (Silver et al. 1979). 

It was also reported at this time that Michigan farm residents with the highest exposure to PBB had 

significantly elevated levels of IgM, IgA, and IgG relative to Wisconsin dairy farm residents (Bekesi et al. 

1985). Cluster analysis of several immunological parameters performed for husbands and wives showed, 

according to the investigators, significant correlations for surface markers, lymphocyte functions, and IgG 

values (no correlation coefficient was >0.337). This finding was interpreted as supporting a common 

dietary source for the immune dysfunction rather than a genetic predisposition (Bekesi et al. 1985). 

In yet another report, Michigan farmers reported a higher rate of infections (11%) than a group of 

chemical workers exposed to PBB (3%) (Stross et al. 1981), however, average PBB levels in serum, bile, 

and fat were higher in the chemical workers than in the farmers. When the patients were divided 

according to their PBB fat level into high, moderate, and low, there was an equal distribution of abnormal 

physical, laboratory, and diagnostic findings among the groups. 
PBBs and PBDEs 196 

5. HEALTH EFFECTS 

The immunological effects of the commercial PBB mixtures FireMaster FF-1 and FireMaster BP-6 have 

been examined in rats, mice, guinea pigs, dogs, and pigs, but in many cases, the most sensitive 

immunological end points were not examined. In all but two studies, the animals were exposed for an 

intermediate duration, and many studies administered the PBBs by gavage (exceptions noted below). 

Additionally, most studies were conducted in rats, a species that may be a poor model for investigating 

dioxin-like effects on the adult immune system. Identification of the most sensitive species is further 

complicated by the fact that not all studies examined the same end points, although limited data suggest 

that guinea pigs may be particularly sensitive. Immunological effects in animals, attributed to exposure to 

PBBs in utero or through lactation, are discussed in Section 5.2.2.6. 

Limited data exist regarding immunological effects of PBBs in animals following acute oral exposure. 

No histopathological alterations were observed in the spleen and thymus of rats treated with a single dose 

of 1,000 mg/kg FireMaster FF-1 and observed for 2 years (Kimbrough et al. 1978b). A similar lack of 

effects in the thymus, spleen, and lymph nodes was reported in rats and mice treated for 2 weeks with up 

to 30 mg/kg/day FireMaster FF-1 (Gupta et al. 1981). However, mice treated with ≈130 mg/kg 

FireMaster BP-6 in the diet for 14 days were incapable of mounting an antibody-mediated response 

following immunization with sheep red blood cells (SRBC) (Fraker 1980; Fraker and Aust 1978). This 

treatment also reduced absolute thymus weight by 88% and caused high lethality in mice. 

Numerous intermediate-duration studies have examined the immunological effects of PBBs in rats. For 

example, treatment of rats with FireMaster FF-1 for 25 weeks significantly increased absolute and relative 

spleen weight at ≥1 mg/kg/day and significantly decreased absolute and relative thymus weight at 

≥0.3 mg/kg/day (NTP 1983). Nevertheless, no histopathological alterations were observed in these 

organs and in lymph nodes with doses of up to 10 mg/kg/day (NTP 1983). Similar results were reported 

in rats treated with 30 mg/kg/day FireMaster FF-1 for 4–5 weeks (Gupta and Moore 1979; Gupta et al. 

1981), but a dose of 100 mg/kg/day caused thymic atrophy and necrosis of lymphoblasts (Gupta and 

Moore 1979). A much smaller dose, 0.5 mg/kg/day FireMaster BP-6 in the diet for 150 days, reportedly 

caused moderate lymphoid depletion in thymus and spleen (Rezabek et al. 1989). In the only chronic rat 

study, splenic fibrosis developed following exposure to 1.5 mg/kg/day, but not 0.5 mg/kg/day, FireMaster 

FF-1 in the diet for up to 104 weeks (NTP 1992). Treatment of rats for 5 weeks with 30 mg/kg/day 

FireMaster FF-1 significantly reduced the in vitro lymphocytic response to stimulation with two out of 

three mitogens and thymus and spleen weight (Luster et al. 1978). Relative thymus weight was reduced 

at 3 mg/kg/day; however, treatment with the test material did not alter the production of antibodies 4 days 

after immunization with SRBC. The same group of investigators reported significantly decreased 
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5. HEALTH EFFECTS 

lymphoproliferative responses to mitogens or allogenic cells in rats following treatment with 3 mg/kg/day 

FireMaster FF-1 for 6 months (Luster et al. 1980); a dose of 1 mg/kg/day was without effect. It must be 

mentioned, however, that in the studies conducted by Luster and co-workers, doses ≥3 mg/kg/day 

FireMaster, reduced body weight by ≥15% in the animals, suggesting that PBBs can affect the immune 

system, but only at dose levels that produce overt toxicity. 

Mice treated for 30 days with FireMaster BP-6 in the diet at levels of approximately ≥1.3 mg/kg/day had 

a significantly reduced antibody-mediated response to SRBC (p<0.001) (Fraker 1980; Fraker and Aust 

1978). Absolute thymus weight was significantly reduced (p<0.01) relative to controls with all dose 

levels tested (0.13, 1.3, 13 mg/kg/day). Delayed-type hypersensitivity was not altered by PBB treatment. 

Corticosterone levels in plasma were elevated in treated mice relative to controls, but not elevated enough 

to be responsible for the immunological findings. No histopathological effects were observed in the 

thymus, spleen, or lymph nodes of mice treated with 30 mg/kg/day FireMaster BP-6 for 4–5 weeks, but 

relative thymus weight was temporarily decreased (Gupta et al. 1981). Other studies in mice reported 

increased lethality (p<0.05) after bacterial inoculation in groups treated with 10 mg/kg/day FireMaster 

FF-1 for 6 months (Luster et al. 1980) and increased lethality (p<0.05) due to challenge with Salmonella 

thyphosa lipopolysaccharide after 3 or 6 weeks of dietary exposure to ≈21.7 mg/kg/day FireMaster FF-1 

(Loose et al. 1981). No histopathological changes were observed in the spleen, thymus, and lymph nodes 

of mice treated with up to 10 mg/kg/day FireMaster FF-1 for 25 weeks (NTP 1983), although 

3.9 mg/kg/day for up to 105 weeks caused increased splenic hematopoiesis (NTP 1992). 

Guinea pigs administered 0.4 mg/kg/day FireMaster BP-6 in the diet for 45 days exhibited a significant 

reduction (p<0.01) in tetanus-antitoxin titers following injection of tetanus toxoid (Vos and van Genderen 

1973, 1974). A dose 5 times higher caused marked thymus atrophy, splenic effects (marked depletion of 

the follicles and periarteriolar lymphocyte sheaths), and lethality. Pregnant sows fed a diet that provided 

approximately 2.5 mg/kg/day FireMaster BP-6 for a total of 12 weeks (including part of gestation and 

lactation) showed a significantly reduced (p<0.05) lymphocyte response to stimulation with PHA and 

PWM mitogens relative to controls (Howard et al. 1980); a dose of 1.25 mg/kg/day was without effect. 

However, PBB treatment did not affect bactericidal activity of whole blood towards Escherichia coli and 

Staphylococcus aureus. 

Two cows gavaged with daily doses of 67 mg/kg/day of an unspecified PBB mixture for 38 consecutive 

days showed minimal alterations in tests of humoral and cell immunity relative to a group of 

54 unexposed animals (Kateley et al. 1982). The concentration of PBBs in tissues from these two cows 
PBBs and PBDEs 198 

5. HEALTH EFFECTS 

reached 1,000 mg/kg, and they became moribund and were later sacrificed. A similar lack of significant 

immunological effects was reported in the same study for 58 cows from contaminated farms in Michigan 

that had PBB body burdens ranging from 0.02 to 24 mg/kg for at least 2 years (Kateley et al. 1982). 

Cows that received gavage doses of ≤0.65 mg/kg/day FireMaster PB-6 for 60 days showed no 

histopathologic alterations in the thymus or spleen (Moorhead et al. 1977). However, doses of 67 mg/kg 

induced thymic involution and atrophy, and were nearly lethal. 

The highest NOAEL values and all LOAEL values from each reliable study for immunological effects in 

each species and duration category are recorded in Table 5-2 and plotted in Figure 5-2. 

Polybrominated Diphenyl Ethers. No studies were located regarding immunological effects in humans 

after oral exposure to PBDEs. No effects on pokeweed mitogen-stimulated DNA proliferation or IgG 

immunoglobulin synthesis were found in human lymphocytes exposed to 2,2',4,4'-tetraBDE (BDE 47) or 

2,2',3,4,4'-pentaBDE (BDE 85) in vitro (Fernlof et al. 1997). 

Limited information is available on effects of acute-duration exposure to pentaBDE on immunologic 

function in animals. A single gavage dose of 0.8–500 mg/kg of pentaBDE (DE-71) did not effect the 

plaque-forming splenic cell (PFC) antibody response to injected SRBC in mice (Fowles et al. 1994). 

Mice that were given 18, 36, or 72 mg/kg/day doses of pentaBDE (DE-71) by gavage for 14 days had 

significantly reduced antibody response to SRBC (63% of control value, p<0.02) and decreased thymus 

weight at 72 mg/kg/day (Fowles et al. 1994). There were no exposure-related effects of the 14-day 

exposure to ≤72 mg/kg/day on natural killer cell (NKC) activity to murine YAC-1 target cells; NKC 

activity was not evaluated in the single dose study. A 14-day study of another pentaBDE mixture 

(Bromkal 70-5 DE) was conducted in which mice and rats were administered 18 or 36 mg/kg/day by 

gavage and were evaluated for spleen and thymus weights, numbers of splenic and thymic lymphocyte 

subsets (CD4+, CD8+, and CD45R+ cells), and in vitro IgG immunoglobulin production in pokeweed 

mitogen-stimulated splenocytes (Darnerud and Thuvander 1998). The only exposure-related effect in 

either species was significantly reduced in vitro production of IgG in pokeweed-stimulated splenocyte 

cultures from the mice exposed to 36 mg/kg/day. The effects in both 14-day studies of commercial 

pentaBDE mixtures occurred only at the highest dose and are possibly due to contamination with PBDDs 

and PBDFs. Mice that were similarly tested with 18 mg/kg/day of the congener 2,2’,4,4’-tetraBDE 

(BDE 47) for 14 days had significantly reduced numbers of total splenocytes as well as CD4+, CD8+, and 

CD45R+ cells in spleen (Darnerud and Thuvander 1998). This was the only dose level of a single 

congener used in the mice. Rats were not evaluated. 
PBBs and PBDEs 199 

5. HEALTH EFFECTS 

Histopathological examinations of spleen, thymus, lymph node and/or bone marrow tissues showed no 

effects of repeated dietary administration in intermediate-duration studies in rats exposed to 

≤8,000 mg/kg/day decaBDE for 13 weeks (NTP 1986), in mice exposed to ≤9,500 mg/kg/day decaBDE 

for 13 weeks, in rats exposed to ≤750 mg/kg/day octaBDE for 13 weeks (IRDC 1977), or in rats exposed 

to ≤100 mg/kg/day pentaBDE for 90 days (WIL Research Laboratories 1984). Chronic ingestion of 

decaBDE caused lesions in the spleen of rats exposed to ≥1,200 mg/kg/day (splenic hematopoiesis in 

males) or 2,240 mg/kg/day (splenic fibrosis and lymphoid hyperplasia in females) for 103 weeks (NTP 

1986). 

The highest NOAEL values and all LOAEL values from each reliable study for immunological effects in 

each species and duration category are recorded in Table 5-3 and plotted in Figure 5-3. 

5.2.2.4 Neurological Effects 

Polybrominated Biphenyls. Neurological symptoms were reported frequently by Michigan residents 

during a 3–4-year period following the 1973 PBB contamination episode, but positive associations 

between serum PBB levels and frequency of neurological symptoms were not found in several studies. In 

an epidemiological study conducted by the U.S. Center for Disease Control and the Michigan Department 

of Public Health, fatigue was reported more frequently by several putatively exposed groups including 

2,148 residents of farms quarantined for PBB contamination (36.4% prevalence rate), 1,421 recipients of 

food from contaminated farms (32.4%), 252 chemical workers involved in PBB manufacturing or 

distribution (22.0%), and 331 residents of farms with low levels of PBB contamination (41.4%), than by a 

small (60 persons) unexposed control group (15.8%); however, no positive association was apparent 

between serum levels of PBB and prevalence rates for any reported symptom (Landrigan et al. 1979). 

Neurological symptoms, including marked tiredness and decrements in the capacity for intellectual and 

physical work, also were reported with greater frequencies in groups of farmers and residents of Michigan 

likely to have consumed farm products contaminated with PBB, than in groups of unexposed Wisconsin 

farmers; however, serum PBB levels were not positively associated with prevalence rates for any 

symptom including neurological symptoms, nor with performance on neurobehavioral tests for a subset of 

this population (Anderson et al. 1978c, 1979; Valciukas et al. 1978, 1979). In a 1976 medical history 

questionnaire study of 342 Michigan children likely to have been exposed to PBBs and 72 unexposed 

children from Wisconsin, the number of subjectively reported symptoms of ill health, including several 

symptoms of neurological effects, did not increase with increasing serum PBB levels (assayed in 1976), 
PBBs and PBDEs 200 

5. HEALTH EFFECTS 

but rather decreased; general neurological examinations did not reveal a pattern of abnormality among the 

Michigan children (Barr 1980). Subjectively reported symptoms of neurological effects including 

weakness, fatigue, difficulty in concentrating, and irritability were prevalent in a group of 23 farmers 

involved in the Michigan PBB contamination episode, but tests of intelligence, memory, and nerve 

conduction velocity failed to demonstrate abnormalities. In addition, a group of 28 workers involved in 

the manufacture or distribution of PBB displayed higher average serum PBB levels than the farmers 

(48 ppb versus 14 ppb), but did not report a prevalence of symptoms of neurological effects (Stross et al. 

1981). In a study of 21 Michigan residents who consumed PBB-contaminated food and had lingering 

medical complaints and 21 volunteer control subjects with putative low-dose exposure to PBB, no 

positive association was observed between PBB levels in fat tissue and performance in a battery of 

neuropsychological tests (Brown and Nixon 1979). In general, the findings of the epidemiological and 

clinical studies of people exposed to PBBs in Michigan are inconclusive; they do not clearly demonstrate 

or eliminate the possibility of an association between PBB oral exposure and the occurrence of 

neurological effects. 

Limited data indicate that orally (gavage) administered PBBs can produce neurological effects in rats. 

FireMaster FF-1 at 10 mg/kg/day (3 days/week) for 8 weeks did not alter the performance of rats in tests 

of operant behavior, but decreased motor activity, grip strength, and startle responsiveness observed in 

rats following administration of ≤10 mg/kg/day for 6 months or 30 mg/kg/day for 4 weeks (Tilson and 

Cabe 1979). Motor activity changes were also observed in rats administered doses of FireMaster FF-1 as 

low as 1 mg/kg/day for 4 weeks (Geller et al. 1979). In this experiment, neither learning nor performance 

of a simple discrimination task was affected by 1, 3, or 6 mg/kg/day dosage levels, but increased motor 

activity was observed at 1 mg/kg/day. No changes were apparent at 3 mg/kg/day and decreased motor 

activity was apparent at 6 mg/kg/day, compared with controls. Weakness of the hind limb was noted in 

rats treated with 10 mg/kg/day FireMaster FF-1 for 6 months compared with control rats (Cabe and 

Tilson 1978). Histological examination of brain and/or spinal nerve tissue found no FireMaster FF-1-

related alterations in rats or mice administered up to 10 mg/kg/day for 25 weeks (NTP 1983) or 

3.9 mg/kg/day for up to 105 weeks (NTP 1992). 

Neurodevelopmental effects were assessed in offspring of mice that were treated with 3 or 10 mg/kg/day 

doses of FireMaster (FF-1) in corn oil by gavage on every other day during gestation and until weaning of 

the offspring at 21 days of age (Tilson 1992). Acoustic startle response, negative geotaxis, motor activity, 

and body weight were measured in 8 pups/sex/dose at 30, 60, and 120 days of age. Tests for avoidance 

learning and neurochemistry were performed on one pup/sex/dose at 30 days of age and on the remaining 
PBBs and PBDEs 201 

5. HEALTH EFFECTS 

animals at 120 days of age. Reductions in acoustic startle responsiveness and negative geotaxis latency 

were observed at 10 and ≥3 mg/kg/day, respectively, in both sexes at 30 and 60 days of age. Motor 

activity was decreased in 10 mg/kg/day females at 120 days of age. The learning tests showed increased 

avoidance response latencies at 30 and 120 days of age in both sexes at ≥3 mg/kg/day, but no effect on 

acquisition or retention. Neurochemical measurements included serotonin and metabolites, dopamine and 

metabolites, and norepinephrine in the cortex, hippocampus, and striation; the only effect observed was a 

decrease in dopamine concentration in the striation of both males and females at 120 days of age. 

Postnatal neurodevelopmental effects were also evaluated in offspring of rats that received 0.2 or 

2 mg/kg/day doses of FireMaster BP-6 dissolved in peanut butter from day 6 of gestation through day 24 

postpartum and observed until postnatal day 60 (Henck et al. 1994). Multivariate analysis of variance of 

neurodevelopmental end points showed significant PBB-related effects for acquisition of forward 

locomotion, cliff avoidance, cage emergence, and open-field activity in male and female offspring of the 

rats exposed to 2 mg/kg/day. The most prominent behavioral effects were delays in acquisition of 

forward locomotion and suppressed open-field activity. Other effects in the offspring included reduced 

crown-rump length and body weight at birth and reduced postnatal body weight as summarized in 

Section 5.2.2.6 (Developmental Effects). 

The highest NOAEL values and all LOAEL values from each reliable study for neurological effects in 

each species and duration category are recorded in Table 5-2 and plotted in Figure 5-2. 

Polybrominated Diphenyl Ethers. No studies were located regarding neurological effects in humans 

after oral exposure to PBDEs. 

A limited amount of information is available on neurological effects of PBDE mixtures in animals. None 

of the commercial decaBDE, octaBDE, and pentaBDE products have been screened for neurotoxicity 

using comprehensive test batteries that typically include functional observational, motor activity, and 

neuropathology evaluations. 

There were no indications of neurotoxicity for commercial decaBDE, as assessed by overt clinical signs 

and nervous system histopathology, in rats and mice in lifetime feeding studies of the 94–97% pure 

mixture (NTP 1986). These animals were exposed to high dietary levels of the decaBDE product 

(estimated doses as high as 2,550 mg/kg/day in rats and 7,780 mg/kg/day in mice) for 103 weeks. There 

also were no overt signs of neurotoxicity in rats and mice exposed to decaBDE in estimated dietary doses 
PBBs and PBDEs 202 

5. HEALTH EFFECTS 

of ≤16,000 and ≤19,000 mg/kg/day, respectively, for 14 days, or ≤8,000 and ≤9,500 mg/kg/day, 

respectively, for 13 weeks (NTP 1986). Histological examinations of the nervous system were not 

performed in these 14-day and 13-week studies. Although the high doses and extended exposure 

durations in the NTP (1986) studies provided opportunities for the induction and/or development of 

effects, neurotoxicity is incompletely evaluated due to the lack of testing for subtle behavioral and other 

sensitive neurological end points. 

Three studies, reported in limited detail as abstracts, assessed neurobehavioral development in offspring 

of rats that were perinatally exposed to commercial pentaBDE mixture DE-71 in corn oil by gavage on 

GD 6 through PND 21, in dose ranges of 1–30 mg/kg/day (Taylor et al. 2002), 1–100 mg/kg/day (Taylor 

et al. 2003), or 5–100 mg/kg/day (MacPhail et al. 2003). Behavioral end points included motor activity 

(MacPhail et al. 2003; Taylor et al. 2002, 2003), auditory startle (Taylor et al. 2002, 2003), and fear 

conditioning (Taylor et al. 2003). Evaluation of adult offspring showed no alterations in motor or sensory 

development as assessed by overall levels of motor activity, within-session habituation of activity, or 

auditory startle response, although tests of fear conditioning revealed decreases in cue-based (but not 

context-based) performance. The lack of motor activity changes in the DE-71 mixture-exposed rats 

differs from attenuating effects on habituation of motor activity observed in mice exposed to individual 

PBDE congeners, as summarized below. 

Neurobehavioral effects of postnatal or perinatal exposure to individual PBDE congeners have been 

evaluated in a number of mouse studies. As detailed below, these studies collectively indicate that the 

developing nervous system is a target of particular congeners, as shown by mild performance 

impairments in tests of spontaneous motor behavior and learning and memory in adult mice. The 

aberrations in spontaneous behavior generally appeared to persist and worsen with age, and were induced 

during a defined critical phase of neonatal brain development. 

In a series of congener studies using similar experimental designs, single doses of 2,2',4,4'-tetraBDE 

(BDE 47), 2,2',4,4',5-pentaBDE (BDE 99), 2,2',4,4',5,5’-hexaBDE (BDE 153), or 2,2',3,3’,4,4',5,5’,6,6’-

decaBDE (BDE 209) were postnatally administered to male NMRI mice in a 20% fat emulsion vehicle by 

gavage (Eriksson et al. 1999, 2001a, 2002a; Viberg et al. 2001a, 2001b, 2002a, 2002b, 2003a). 

Evaluations included spontaneous motor behavior at 2, 4, and/or 6 months of age (BDEs 47, 99, 153, and 

209), and Morris swim maze performance at 5 or 6 months of age (BDEs 47, 99, and 153). The 

spontaneous behavior test measured open-field locomotion (horizontal movements), rearing (vertical 

movements), and total activity (all types of vibrations within a test cage). The measurements were 
PBBs and PBDEs 203 

5. HEALTH EFFECTS 

performed during three consecutive 20-minute periods to assess habituation, defined as a decrease in the 

motor activities in response to diminishing novelty of the test chamber over 60 minutes. The water maze 

test assessed spatial learning ability and memory by evaluating the ability to locate a submerged platform. 

Cognition was tested by providing visual cues to find and remember the location of a platform submerged 

in a pool of water. The decrease in time needed to locate the platform over a total of 20 trials in a 4-day 

acquisition period was used as a measure of learning ability. On day 5, the platform was relocated and 

the decrease in time needed to find the relocated platform over five trials was used to assess relearning 

ability. 

In the study with 2,2',3,3’,4,4',5,5’,6,6’-decaBDE (BDE 209), the mice were dosed once with 0, 2.22, or 

20.1 mg/kg on PNDs 3 or 19, or 0, 1.34, 13.4 or 20.1 mg/kg on PND 10, and evaluated for spontaneous 

behavior at 2, 4, and 6 months of age (Viberg et al. 2001b, 2003a). A non-habituating behavior profile 

was observed in the mice exposed on PND 3, but not in those exposed on PND 10 or 19. In particular, 

following exposure on PND 3, aberrations in spontaneous motor behavior were manifested at 6 months of 

age in the mice exposed to 2.22 mg/kg, and at 2, 4, and 6 months of age in those exposed to 20.1 mg/kg. 

The aberrations were more pronounced in the older animals, indicating that the effects worsened with age. 

The findings are consistent with tissue distribution data in this study, suggesting that brain uptake of 

14C-BDE 209 was more efficient in the younger animals and that the amount of radioactivity reaching the 

brain increased with time (see Section 5.4.2). Additional information on the behavioral results of this 

study, including quantitative data, was not reported. 

In the study with 2,2',4,4',5,5’-hexaBDE (BDE 153), mice were dosed once with 0, 0.45, 0.9, or 

9.0 mg/kg on PND 10, and evaluated for spontaneous behavior at 2, 4, and 6 months of age, and swim 

maze performance at 6 months of age (Viberg et al. 2001b, 2002a). Non-habituating behavior occurred in 

mice exposed to 0.45 mg/kg at 6 months of age, and 0.9 and 9.0 mg/kg at 2, 4, and 6 months of age. The 

effect worsened with age and was most pronounced at 9.0 mg/kg. The swim maze test found that all mice 

improved their ability to locate the platform during the acquisition period, although those exposed to 

≥0.9 mg/kg had significantly longer latencies to locate the platform on days 2 and 3 during the acquisition 

period. Additional information on the behavioral results of this study, including quantitative data, was 

not reported. 

Mice were also exposed once to 2,2',4,4'-tetraBDE (BDE 47) in doses of 0, 0.7, or 10.5 mg/kg, or 

2,2',4,4',5-pentaBDE (BDE 99) in doses of 0, 0.8, or 12.0 mg/kg, on PND 10 (Eriksson et al. 1998, 

2002b). Spontaneous behavior was tested at 2 and 4 months of age in all dose groups, and swim maze 
PBBs and PBDEs 204 

5. HEALTH EFFECTS 

performance was tested at 5 months of age in the groups given the high dose of each congener. A 

disruption of habituation was observed in mice exposed to both congeners. During the first of the three 

consecutive 20-minute test periods, at both 2 and 4 months, animals treated with 10.5 mg/kg BDE 47 and 

≥0.8 mg/kg BDE 99 were significantly less active than controls as shown by dose-related decreases in all 

three activity measures. During the second period, the activities in all treated groups were comparable to 

controls at 2 and 4 months. During the third period, exposure to 10.5 mg/kg BDE 47 and ≥0.8 mg/kg 

BDE 99 caused significantly more activity than controls at 2 and 4 months as shown by dose-related 

increases in the test measures. The investigators noted that this nonhabituating behavior profile (i.e., 

decreased activity early in the test period and increased activity late in the test period) has also been 

reported in adult mice neonatally exposed to certain PCB congeners. The same non-habituating behavior 

at 2 and 4 months of age (hypoactivity followed by hyperactivity) seen in mice exposed to an 8 mg/kg 

dose of BDE 99 on PND 10 (Viberg et al. 2002b). The only exposure-related effect in the Morris water 

maze test was found in the mice exposed to BDE 99 at 12.0 mg/kg; performance in finding a new 

platform location did not improve as it did in controls and mice exposed to BDE 47(Eriksson et al. 1998, 

2001a). 

Other reports further characterized effects of postnatal exposure to 2,2’,4,4’,5-pentaBDE (BDE 99) on 

spontaneous motor behavior in mice. In a study designed to ascertain if NMRI male mice are susceptible 

during a critical phase of neonatal brain development (i.e., during the period of rapid brain growth), 

spontaneous behavior was evaluated in adult mice that were administered a single 8 mg/kg dose of BDE 

992 on PNDs 3, 10, or 19 and tested at 4 months of age (Eriksson et al. 1999, 2002b). A non-habituating 

behavioral pattern similar to that observed in the studies summarized above (i.e., decreased locomotion, 

rearing, and total activity over time in response to diminished novelty of the test chambers) was seen in 

the mice treated at either 3 or 10 days of age. There were no significant changes in the three spontaneous 

activity measures in the mice exposed at 19 days of age, suggesting that there was a critical window for 

the induction of the behavioral disturbances. In an assessment of sex and/or strain dependency (i.e., 

whether mice other than NMRI males were affected), male and female C57 Bl/J mice were exposed to 

BDE 99 as a single gavage dose of 0.4, 0.8, 4.0, 8.0, or 16.0 mg/kg on PND 10, and tested at 2, 5, and 

8 months of age (Viberg et al. 2003a). Spontaneous motor behavior was significantly impaired at 

≥0.8 mg/kg in both sexes at all three ages. The effect was dose-related and appeared to worsen with age, 

showing a pattern of response similar to that observed for BDE 99 and other congeners in the various 

studies with NMRI male mice. 
PBBs and PBDEs 205 

5. HEALTH EFFECTS 

In a study of perinatal exposure to 2,2’,4,4’,5-pentaBDE (BDE 99), groups of two male and two female 

CD-1 Swiss female mice were gavaged with 0, 0.6, 6, or 30 mg/kg/day in corn oil from GD 6 to PND 21 

(Branchi et al. 2001, 2002). Somatic development (body weight gain, hair growth, day of eyelid and ear 

opening, day of incisor eruption) and neurobehavioral development (righting reflex, forelimb stick 

grasping reflex, forelimb placing reflexes, negative geotaxis, screen grasping and climbing, pole grasping, 

ultrasonic vocalizations, homing test) were assessed during PNDs 2–20. Spontaneous activity, including 

locomotion (horizontal movement), rearing (vertical movement), and thigmotaxis (time and distance 

traveled close to walls, an index of emotionality), was assessed on PNDs 22, 34, 60, and 120. There were 

no clear exposure-related indications of prenatal toxicity as shown by maternal weight gain and 

pregnancy and litter indices. The testing during PNDs 2–22 suggested a delay in sensorimotor 

development at 30 mg/kg/day, as indicated by approximately 2 days delayed maturation in screen 

climbing response. The spontaneous behavior testing showed changes suggestive of an age-dependent 

alteration in activity at ≥6 mg/kg/day; effects included hyperactivity (increased locomotion and rearing) 

and impaired habituation at PNDs 34 and 60, altered thigmotaxis (reduced time near walls) at PND 60, 

and a tendency to hypoactivity (reduced locomotion) at PND 120. 

5.2.2.5 Reproductive Effects 

Polybrominated Biphenyls. Analysis of semen from 41 male residents of Michigan who lived on PBBcontaminated farms or who had bought food directly from such farms and 11 males who were employed 

in a PBB manufacturing company revealed no abnormalities in the distribution of sperm counts, sperm 

motility, or sperm morphology, compared with an analysis of semen from 52 unexposed men (Rosenman 

et al. 1979). This study was conducted in 1977, ≈4 years after initial contamination of Michigan's food 

supply, and would not have detected an earlier response that was subsequently reversed. PBBs were 

detected (detection limit of 0.2 ppb) in the serum of 1 of the 52 unexposed men and in all of the exposed 

men; however, individual or mean values for PBB levels were not reported. 

No relationship was found between serum levels of PBBs and frequency and duration of lactation in a 

retrospective study of women exposed to PBBs during the Michigan contamination episode (Thomas et 

al. 2001). A group of 1,020 women with available initial serum PBB levels was identified from the 

Michigan Department of Community Health PBB registry. Among these participants, 446 had a liveborn infant after their initial serum PBB level; characteristics of this cohort included mean age of 

25.6±5.0 years, initial serum PBB level of 17.5±99.7 ppb, estimated serum PBB level at delivery of 

9.4±61.9 ppb, estimated serum PCB level at delivery of 5.5±5.2 ppb, duration of breast-feeding as main 
PBBs and PBDEs 206 

5. HEALTH EFFECTS 

source of nutrition of 2.6±3.3 months, and total duration of breast feeding of 4.1±5.3 months. The 

numbers of women who breast fed their first infant after the initial serum PBB level and had previously 

breast-fed were 293 (65.7%) and 49 (11.0%), respectively. Exposure was treated as a categorical variable 

by dividing the women into groups of low (reference) exposure (≤1 ppb, n=260, serum levels at or below 

the detection limit), moderate exposure (>1–≤7 ppb, n=141), and high exposure (>7 ppb, n=45, levels 

above the 90th percentile). Three outcomes of interest were analyzed: (1) the decision to breast feed 

(yes/no), (2) the duration (months) of breast-feeding as the main source of nutrition, and (3) the total 

duration (months) of breast-feeding. None of the three outcomes was significantly associated with serum 

PBB levels, even after controlling for maternal age, previous history of breast-feeding, body mass index, 

maternal education, household income, history of smoking in the year before pregnancy, consumption of 

alcohol during the first trimester of pregnancy, history of thyroid disorder, gestational age of the infant, 

time to pregnancy, and year of birth. 

Effects on reproductive organs and reproductive function have been observed in animals following oral 

exposure to PBBs. An increased incidence of uterine endometrial polyps was observed in rats, 2 years 

after they were administered a single gavage of 1,000 mg/kg dose FireMaster FF-1 (Kimbrough et al. 

1981). Following weaning and two consecutive normal menstrual cycles in 6 months, serum 

progesterone was decreased in the same four females that showed this effect prebreeding. In a multiplegeneration study in which only F0 rats were fed ≥5 mg/kg/day FireMaster BP-6 in the diet from 

GD 8 through postpartum day 28 (weaning), reproductive performance with respect to length of gestation 

or litter size was not affected in the F1 or F2 generations (McCormack et al. 1981). Implantation was 

completely blocked in two of five and two of three female rats that survived gavage administration of 

28.6 or 57.1 mg/kg/day FireMaster BP-6, respectively, on alternate days between GDs 0 and 

14 (Beaudoin 1979). Histological examination of reproductive organs in male and female rats and mice 

revealed no abnormalities following gavage treatment with doses up to 10 mg/kg/day FireMaster FF-1 for 

25 weeks or 30 mg/kg/day for 4–5 weeks (Gupta and Moore 1979; Gupta et al. 1981; NTP 1983). 

Necrosis, hyperplasia, and metaplasia in the epithelial lining of the ductus deferens were observed in male 

rats that died following 100 mg/kg/day FireMaster FF-1 for 4–5 weeks (Gupta and Moore 1979). 

Treatment of male and female mink with diets providing up to 0.39 mg/kg/day FireMaster FF-1 for 6–

7 months before breeding did not affect reproductive performance with respect to fertility or litter size 

(Aulerich and Ringer 1979; Ringer et al. 1981). In the only chronic study, histological examination of 

male and female reproductive organs showed increased cystic endometrial hyperplasia in rats exposed to 

1.5 mg/kg/day FireMaster FF-1 for up to 104 weeks, but no changes were observed in mice exposed to 

≤3.9 mg/kg/day for up to 105 weeks (NTP 1992). Following 6–7 months of exposure to 0.012 mg/kg/day 
PBBs and PBDEs 207 

5. HEALTH EFFECTS 

FireMaster FF-1 in the diet, four of seven female monkeys displayed a lengthening of the menstrual cycle 

from 28 to 31 days and decreased serum progesterone; prior to the treatment, they had at least 2 years of 

normal cycles (Allen et al. 1979; Lambrecht et al. 1978). All seven of these monkeys conceived after one 

to four matings with control males (controls required one to three breedings), but two displayed prolonged 

implantation bleeding and another two had fetuses that were aborted or stillborn (see Developmental 

Effects). Reduced spermatogenesis was observed in a male monkey that died after 25 weeks on a diet 

providing 0.73 mg/kg/day FireMaster FF-1 (Allen et al. 1978). 

The highest NOAEL values and all LOAEL values from each reliable study for reproductive effects in 

each species and duration category are recorded in Table 5-2 and plotted in Figure 5-2. 

Polybrominated Diphenyl Ethers. No studies were located regarding reproductive effects in humans 

after oral exposure to PBDEs. 

Information on effects of PBDEs on reproductive function is limited to negative findings in a onegeneration study of decaBDE (FR-300 BA) in rats (Dow Chemical Co. 1975; Norris et al. 1975b). This 

commercial mixture was comprised of 77.4% decaBDE, 21.8% nonaBDE, and 0.8% octaBDE and differs 

from current decaBDE formulations that contain ≥97% decaBDE. Male and female rats were exposed to 

3, 30, or 100 mg/kg/day doses in the diet for 60 days before mating and subsequently during a 15-day 

period during which they were mated. Both sexes continued to receive the test diet throughout gestation 

and until the end of the 21-day lactation period. Parameters monitored included length of time between 

first day of cohabitation and parturition, numbers of live and dead newborn, number of live pups (PNDs 

1, 7, 14, and 21), litter weight (PNDs 1, 7, and 14), and weanling weight (PND 21). Comprehensive 

histological examinations (adults and weanlings), skeletal examinations (weanlings), and cytogenetic 

evaluation of bone marrow (adults and weanlings) were also performed on PND 21. There were no 

exposure-related effects on reproductive parameters or any indications of maternal or neonatal toxicity. 

No histopathological changes were observed in male or female reproductive tissues from rats that were 

exposed to decaBDE in dietary doses of ≤800 mg/kg/day for 30 days (Norris et al. 1973, 1975b), 

≤8,000 mg/kg/day for 13 weeks (NTP 1986), ≤1.0 mg/kg/day (77.4% containing 21.8% nonaBDE and 

0.8% octaDBE) for 2 years (Kociba et al. 1975; Norris et al. 1975b), or ≤2,550 mg/kg/day for 103 weeks 

(NTP 1986), or in mice exposed to ≤9,500 mg/kg/day for 13 weeks or ≤7,780 mg/kg/day for 103 weeks 

(NTP 1986); octaBDE in doses of ≤750 mg/kg/day for 13 weeks (IRDC 1977); or pentaBDE in doses of 

≤100 mg/kg/day for 90 days (WIL Research Laboratories 1984). 
PBBs and PBDEs 208 

5. HEALTH EFFECTS 

Exposure to lower brominated PBDEs has been reported to cause reproductive effects in developing rats. 

As summarized in Section 5.2.2.6 Developmental Effects, postnatal exposure to the commercial 

pentaBDE mixture DE-71 caused delayed onset of puberty in male and female rats (Laws et al. 2003; 

Stoker et al. 2003). Additionally, testicular and ovarian effects occurred in male and female adult 

offspring of rats that were gestationally exposed the pentaBDE congener BDE 99 (Kuriyama and 

Chahoud 2003; Talsness et al. 2003). 

The highest NOAEL values and all LOAEL values from each reliable study for reproductive effects in 

each species and duration category are recorded in Table 5-3 and plotted in Figure 5-3. 

5.2.2.6 Developmental Effects 

Polybrominated Biphenyls. Examination of children (≈100 were identified) presumably exposed in utero

or in early infancy during the peak of the Michigan PBB contamination episode and whose families lived 

on farms known to be contaminated with PBBs has not revealed any consistent or marked abnormalities 

in the children's physical and neuropsychological development. No significant abnormalities were found 

by physical and neurological examination of 33 of these exposed children when they had a mean age of 

37.2 months, compared with a group of 20 age-matched, nonexposed control children (Weil et al. 1981). 

However, subjective interviews with parents suggested that more exposed children than control children 

had frequent upper respiratory illnesses such as colds, runny noses, and sore throats (Weil et al. 1981). 

PBBs were measured in the fat of the infants and in the blood of the mothers. Fat levels of PBBs in 27 of 

the children ranged from 0.01 to 20.96 ppm; half of the values were below 0.120 ppm, and five of the 

values were above 1.0 ppm. Maternal blood levels ranged from 0.001 to 0.845 ppm and seven mothers 

had levels that were not detectable (<0.001 ppm). Seagull (1983) administered 5 of 18 tests in a battery 

of childhood developmental tests (McCarthy Scales of Children's Abilities) to 19 of these exposed 

children when their ages ranged from ≈2.5–4 years old and concluded that there was a statistically 

significant negative correlation for four of the five tests between PBB levels in fat tissue and 

developmental abilities. Mean fat concentrations of PBBs were 0.50 ppm (range, 0.10–0.74 ppm) and 

4.218 ppm (range, 0.116–20.960 ppm) in the low and high exposure groups of this study. Schwartz and 

Rae (1983) later administered the full battery of neuropsychological developmental tests, as well as I.Q. 

tests, to the same group of children (minus one child whose family refused to participate in the follow-up 

study) when their ages ranged from approximately 4 to 6 years old. The exposed children's performances 

were within the normal range in all areas assessed. There were statistically significant negative 
PBBs and PBDEs 209 

5. HEALTH EFFECTS 

correlations between PBB levels in adipose tissue (measured in the earlier study) and performance on 

some of the developmental tasks, but the tasks with significant correlations were not the same as those 

noted in the earlier study by Seagull (1983). The available studies, primarily due to the small data set and 

the inconsistency of the results, do not conclusively establish or eliminate the possibility that in utero and 

early infancy exposure to PBBs might adversely affect the development of human children. The 

information suggests that if the Michigan PBB contamination episode produced any effects on child 

development, they were subtle. 

A comparison of 1966–1981 fetal mortality rates for Michigan counties with a high percentage of 

quarantined farms (6.8–20.4%) with those of Michigan counties with no quarantined farms did not 

conclusively establish differences in rates or trends after the 1973 contamination episode (Humble and 

Speizer 1984). This study is difficult to interpret because the exposure status method was imprecise, the 

collected data included only spontaneous abortions occurring after 20 weeks of gestation (early trimester 

abortions were not counted), and the two populations displayed different pre-1973 trends for fetal 

mortality rates. 

Results from animal studies indicate that in utero exposure to PBBs and exposure to PBBs through 

mothers' milk can produce embryolethal effects, structural abnormalities, growth retardation, liver effects, 

and neurological effects in offspring. Developmental toxicity has been observed in studies with 

hexabromobiphenyl and octabromobiphenyl commercial mixtures, but not with commercial 

decabromobiphenyl PBBs. Rats have been studied most extensively, but data are also available for mice, 

swine, minks, and monkeys. 

Following gavage administration of 200 mg/kg FireMaster FF-1 to rats on gestation days 7 and 14, 

decreased pup survival to weaning, decreased body weight throughout the lives of offspring, and 

increased mortality in offspring after 2 years were observed (Groce and Kimbrough 1984). Single doses 

of 200 mg/kg FireMaster BP-6 administered to rats on one of several days during pregnancy caused 

increased fetal resorptions, and 400 or 800 mg/kg produced maternal toxicity (expressed as a decrease in 

body weight gain) and fetal malformations including cleft palate and diaphragmatic hernia (Beaudoin 

1977). Increased fetal resorptions also were observed in rats receiving total doses of ≈14.3 mg/kg/day 

FireMaster BP-6 by gavage on alternate days from days 0 through 14 of pregnancy (Beaudoin 1979). 

Body weight gain and levels of vitamin A in the liver were reduced in offspring of rats administered 

5 mg/kg/day FireMaster BP-6 in the diet on gestation day 8 until weaning at 4 weeks postpartum 

(McCormack et al. 1982c). Additional effects in pups weaned onto the same treated diets as the dams 
PBBs and PBDEs 210 

5. HEALTH EFFECTS 

included increased hepatic ALA synthetase activity (the rate-limiting enzyme in porphyrin synthesis) and 

increased urinary excretion of uro- and coproporphyrins at age 16 weeks. Dietary administration of 

FireMaster BP-6 in dosages of 42.9 mg/kg/day on GDs 7–15 or 50 mg/kg/day on GDs 7–20 produced 

decreased body weight, but no other developmental effects, in rat fetuses and pups monitored up to 

60 days postpartum (Corbett et al. 1975; Harris et al. 1978b). Increased incidences of fetuses with extra 

ribs were found in rats fed diets providing ≥86 mg/kg/day of octabromobiphenyl mixture from gestation 

days 6 through 15 (Waritz et al. 1977); however, no embryotoxic, fetotoxic, or teratogenic effects 

occurred in rats following gavage administration of ≤1,000 mg/kg/day decabromobiphenyl mixture on 

GDs 6–15 (Millischer et al. 1980). 

Effects in offspring of rats exposed to 0.5 mg/kg/day FireMaster FF-1 for 60 days before breeding until 

8 weeks postpartum (4 weeks postweaning) and observed for up to the following 2 years included 

vacuolization and altered foci in the liver (Chhabra et al. 1993; NTP 1992). Pups of mice that were 

similarly perinatally exposed to 1.5 mg/kg/day FireMaster FF-1 developed liver cytomegaly and altered 

foci (Chhabra et al. 1993; NTP 1992). As discussed in Section 5.2.2.8 (Carcinogenic Effects), these mice 

also developed hepatocellular adenoma and carcinoma; combined perinatal and adult exposure induced 

higher incidences of liver tumors in mice than adult exposure alone (Chhabra et al. 1993; NTP 1992). 

In a multiple-generation study, decreased pup survival to weaning, decreased body weight gain, delayed 

fur development, delayed eye and vaginal opening, and increased liver weight associated with hepatocyte 

swelling, vacuolization, and focal necrosis were observed in F1 generation rats whose only exposure was 

from the mothers fed a diet providing 5 mg/kg/day FireMaster FF-1 from day 8 of pregnancy until 

weaning at 28 days postpartum; less severe liver responses were observed in the F1 offspring of dams 

treated with 0.5 mg/kg/day (McCormack et al. 1981). Although survival of F2 and F3 generations was not 

affected by the 5 mg/kg/day treatment of the F0 rat dams, F2 offspring, but not F3 offspring, displayed 

increased liver weights, liver enzyme induction, and hepatic histological alterations compared with 

controls (McCormack et al. 1981). Dietary administration of 2.5 mg/kg/day FireMaster BP-6 to rats from 

gestation day 0 through postpartum day 15 produced increased relative liver weights, decreased body 

weights, and decreased serum levels of the thyroid hormone, T4, in 15-day-old offspring (Meserve et al. 

1992). The pups had received direct stimulation of the pituitary by an injection of corticotropin-releasing 

factor or direct stimulation of the adrenals by an injection of adrenocorticotropic hormone. Provision of a 

diet containing 0.5 mg/kg/day FireMaster FF-1 to lactating rats for the 18 days following parturition 

increased liver weights and elevated levels of hepatic cytochrome P-450 and associated enzymic activities 

in both dams and pups; a diet providing 0.05 mg/kg/day produced no hepatic effects in dams, but induced 
PBBs and PBDEs 211 

5. HEALTH EFFECTS 

hepatic enzymes in the pups (Moore et al. 1978). According to the investigators (Moore et al. 1978), the 

results could indicate that nursing pups are more sensitive than their dams to liver enzyme induction, or 

that due to the different kinetic parameters among the PBB congeners, the pups received a more potent 

PBB mixture than the dams. Yet, a third possibility is that the suckling pups received a higher dose of 

PBBs relative to their body weights due to bioconcentration of PBBs in breast milk (Dent 1978). 

Performance deficits in tests of operant behavior were observed in the 6-month-old offspring of rat dams 

given gavage doses of 0.2 or 2 mg/kg/day FireMaster BP-6 from day 6 of gestation until weaning (Henck 

and Rech 1986), but not in 75-day-old offspring of rat dams given gavage doses of 0.5 or 5 mg/kg/day for 

4 weeks prior to mating (Geller et al. 1985). Effects found in offspring of rats exposed to 0.2 or 

2 mg/kg/day doses of FireMaster BP-6 in the diet from day 6 of gestation through day 24 postpartum, and 

observed through postnatal day 60, included reduced crown-rump length at birth at ≥0.2 mg/kg/day, 

reduced birth weight and postnatal body weight gain at 2 mg/kg/day, and suppressed acquisition of 

forward locomotion, cliff avoidance, cage emergence, and open-field activity at 2 mg/kg/day (Henck et al. 

1994). 

Dietary administration of 50 mg/kg/day FireMaster BP-6 to mice from gestation days 7 through 

18 produced decreased fetal body weight and fetal abnormalities including exencephaly, cleft palate, and 

hydronephrosis; 5 mg/kg/day did not produce significant developmental effects in this study (Corbett et 

al. 1975). Early postnatal deaths occurred among offspring of mice given 10 mg/kg/day FireMaster FF-1 

on alternate days from gestation day 0 until litters were weaned (Luster et al. 1980). Immunological 

parameters were unaffected in surviving offspring whose mothers received up to 10 mg/kg/day doses, but 

decreased hematocrit levels were measured in offspring of mothers receiving doses ≥3 mg/kg/day (Luster 

et al. 1980). Performance deficits in tests of learning behavior were measured in offspring of female mice 

that received gavage doses of 3 or 10 mg/kg/day FireMaster FF-1 on alternate days from gestation 

day 0 through weaning at 21 days of age (Tilson 1992). 

Decreased body weight at birth and at 4 weeks after birth were measured in mink kits whose parents were 

fed diets containing 0.155 mg/kg/day FireMaster FF-1 from 7–8 months prior to mating through 4 weeks 

postpartum (Aulerich and Ringer 1979; Ringer et al. 1981). Increased relative liver weight, fatty and 

necrotic hepatic changes, slight hyperplasia in the thyroid, and decreased serum levels of thyroid T3 and 

T4 hormones were observed in 4-week-old offspring of swine fed 2.5 mg/kg/day FireMaster BP-6 in the 

diet during the second half of gestation and during lactation; 1.25 mg/kg/day produced similar effects on 

the thyroid, but no necrosis in the liver of 4-week-old nursing pigs (Werner and Sleight 1981). 

Examination of several parameters of immune function in 4-week-old offspring of sows fed 
PBBs and PBDEs 212 

5. HEALTH EFFECTS 

≤2.5 mg/kg/day FireMaster BP-6 during gestation and lactation provided no conclusive evidence for 

immunosuppressive effects (Howard et al. 1980). An abortion and a stillbirth occurred among seven 

female monkeys that were fed 0.012 mg/kg/day FireMaster FF-1 in the diet for 7 months prior to 

conception and during pregnancy (Allen et al. 1979; Lambrecht et al. 1978). The surviving five infants 

had reduced birth weight and postnatal body weight gain, but no gross abnormalities. The incidence of 

dystocia (difficult birthing) was 50% increased among first- and second-generation offspring of cows 

treated with 0.65 mg/kg/day FireMaster BP-6 by gelatin capsule for 180 or 202 days during late 

pregnancy (Willett et al. 1982). The same dosage for 60 days caused a 21.6% increased incidence of 

dystocia, but this increase was not statistically significant (p=0.08). Stillbirths and preweaning deaths 

were not significantly increased, but all mortality was attributable to dystocia. Incidences of dystocia and 

calf mortality appeared to be related to higher birth weight, which in turn were correlated with 

concentrations of PBBs in maternal blood and tissues. Growth and development were not affected in the 

surviving calves. Of six pregnant cows that were similarly treated with a maternolethal dosage 

(67 mg/kg/day) of FireMaster BP-6, three aborted after 30–38 days, and three retained dead fetuses 

(Moorhead et al. 1977). 

The highest NOAEL values and all LOAEL values from each reliable study for developmental effects in 

each species and duration category are recorded in Table 5-2 and plotted in Figure 5-2. 

Polybrominated Diphenyl Ethers. No studies were located regarding developmental effects in humans 

after oral exposure to PBDEs. 

Information on the developmental toxicity of PBDEs is available from studies of commercial mixtures of 

deca-, octa- and pentaBDE (Argus Research Laboratories 1985b; Breslin et al. 1989; Dow Chemical Co. 

1975, 1985; Hardy et al. 2001, 2002; Life Science Research Israel Ltd. 1987; Norris et al. 1975b; WIL 

Research Laboratories 1986). None of the commercial BDE mixtures have been shown to be overtly 

teratogenic in animals, although neurobehavioral tests, summarized in Section 5.2.2.4 (Neurological 

Effects), indicate that the developing nervous system is a potential target of some individual lower 

brominated congeners, including 2,2’,4,4’-tetraBDE (BDE 47), 2,2’,4,4’,5-pentaBDE (BDE 99), and 

2,2’,4,4’,5,5’-hexaBDE (BDE 153). 

No prenatal developmental toxicity was found in a comprehensive study of commercial decaBDE 

product (Hardy et al. 2001, 2002). The test material was a composite of current commercial decaBDE 

mixtures produced by three manufacturers and had a composition of 97.34% decaBDE and 2.66% nona- 
PBBs and PBDEs 213 

5. HEALTH EFFECTS 

and octaBDE congeners. The mixture was administered to groups of 25 mated female Sprague-Dawley 

rats by gavage in corn oil in daily doses of 0, 100, 300, or 1,000 mg/kg/day on GDs 0 through 19. All 

females were necropsied following sacrifice on GD 20. Maternal end points included clinical 

observations, body weight and food consumption, gravid uterine and liver weights, gross lesions, and 

uterine implantation indices (e.g., numbers of corpora lutea, implantations, and early and late resorptions). 

Fetal end points included viability, body weights, sex distribution, gross malformations (all fetuses), 

skeletal/cartilaginous malformations and ossification variations (approximately half of each litter), and 

visceral malformations (remaining fetuses). There were no treatment-related effects on any of the 

maternal or fetal end points, indicating that 1,000 mg/kg/day was a NOAEL for developmental toxicity of 

decaDBE. 

A lower purity commercial decaBDE product (77% decaDBE, 22% nonaBDE, 0.8% octaBDE) used in 

the 1970s was fetotoxic in rats at high dose levels that were not maternally toxic. Developmental effects 

were investigated in rats that were exposed to doses of 10, 100, or 1,000 mg/kg/day by gavage on GDs 6–

15 and examined on GD 21 (Dow Chemical Co. 1985; Norris et al. 1975b). No treatment-related 

maternal toxicity was observed. The numbers of fetuses with subcutaneous edema and delayed 

ossification of normally developed skull bones were significantly increased at 1,000 mg/kg/day. 

Resorptions were significantly (p<0.05) increased at ≥10 mg/kg/day compared to controls as indicated by 

resorption/implantation site percentages [1% (3/288), 9% (12/141), 10% (13/135), and 4% (9/203)] and 

percentages of litters with resorptions [12% (3/25), 64% (9/14), 57% (8/14), and 39% (7/18)]. The 

resorptions were considered secondary to unusually low control values and unrelated to treatment because 

(1) the data do not follow a dose-response relationship across the three dose levels, and (2) the rates in the 

high dose group are comparable to historical control values. As discussed in Section 5.2.2.5 

(Reproductive Effects), a one-generation study of the 77% commercial decaBDE mixture in rats found no 

effects of parental exposure to ≤100 mg/kg/day from 60 days before mating through the end of the 

lactation on numbers of live pups at birth and during the lactation period, body weights of pups at birth or 

weaning, or skeletal development or soft-tissue histology of pups at weaning (Dow Chemical Co. 1975; 

Norris et al. 1975b). 

Developmental toxicity testing of another octaBDE mixture (DE-79) in rats used gavage doses of 2.5, 10, 

15, 25, or 50 mg/kg/day on GDs 6–15 (WIL Research Laboratories 1986). No exposure-related maternal 

or overt developmental effects were observed at ≤15 mg/kg/day. The only statistically significant 

(p<0.05) finding at 25 mg/kg/day was an increased serum bromide level. Effects observed at 

50 mg/kg/day included significantly reduced mean maternal body weight gain during the posttreatment 
PBBs and PBDEs 214 

5. HEALTH EFFECTS 

period (GDs 16–20) and fetotoxicity as indicated by increased postimplantation loss due to late 

resorptions (not significantly increased compared to control group but exceeded historical control range), 

39% reduced mean fetal weight (p<0.01), skeletal variations (e.g., reduced ossification of the skull and 

various unossified bones) that were associated with the reduced fetal weights in this group, and single 

instances of malformations (fetal anasarca, bent limb bones, unilateral absence of 13th rib) commonly 

associated with maternal toxicity. 

The developmental toxicity of a fourth octaBDE commercial mixture (Saytex 111) was tested in rabbits 

exposed to doses of 2, 5, or 15 mg/kg/day on GDs 7–19 and examined on GD 28 (Breslin et al. 1989). 

The 15 mg/kg/day group showed evidence of slight maternal toxicity as indicated by decreased body 

weight gain during GDs 7–20 and 7–28 (not statistically identified), reduced body weight on GD 28 (7% 

less than controls, p≤0.05), and significantly increased absolute and relative liver weights on GD 28. 

Slight fetotoxicity accompanied the maternal toxicity at 15 mg/kg/day as indicated by a significantly 

(p≤0.05) increased incidence of delayed ossification of the sternebrae. 

A developmental toxicity study of pentaBDE (Saytex 115) is available in which rats were administered 

10, 100, or 200 mg/kg/day doses by gavage on GDs 6–15 and examined on GD 20 (Argus Research 

Laboratories 1985a). The only exposure-related maternal effect was significantly (p≤0.01) reduced body 

weight gain at ≥100 mg/kg/day during the dosing period. This effect was dose-dependent and increased 

in severity with continued dosing but recovered during the postdosage period when differences between 

exposed and control groups were no longer significant. Maternal body weight gain was 2.8, 20.1, and 

29.9% less than controls during GDs 6–16 at 10, 100, and 200 mg/kg/day, respectively. There were no 

exposure-related indications of developmental toxicity. Average fetal body weight per litter was slightly 

but not significantly reduced at 200 mg/kg/day (2.6% less than controls, p>0.05). 

Developmental toxicity was also evaluated in rats that were treated with 1, 10, or 30 mg/kg/day of 

pentaBDE (DE-71) in corn oil by gavage from GD 6 through PND 21 (Zhou et al. 2002). Dams were 

sacrificed on GD 20 and PND 22 and offspring were sacrificed on GD 20 and PNDs 4, 14, 36, and 90. 

There were no exposure-related effects on maternal body weight gain, litter size, sex ratio, or offspring 

viability and growth as assessed by numbers of pups at birth and on PNDs 4–21, body weight of pups on 

PNDs 4–90, and eye opening status on PNDs 11–18. Serum measurements of thyroid T3 and T4 hormone 

levels showed that serum T4 was significantly reduced in the rat dams at 30 mg/kg/day (GD 20 and 

PND 22) and offspring at ≥10 mg/kg/day (GD 20 and PNDs 4 and 14), as detailed in Endocrine Effects in 

Section 5.2.2.2. There were no clear effects on pregnancy and birth indices in mice that were gavaged 
PBBs and PBDEs 215 

5. HEALTH EFFECTS 

with the congener 2,2’,4,4’,5-pentaBDE (BDE 99) in oral doses of 0.6–30 mg/kg/day from GD 6 to 

PND 21 (Branchi et al. 2001, 2002). 

A limited amount of information is available on neurodevelopmental effects of commercial pentaDBE 

mixture DE-71 in mice in two studies incompletely reported as abstracts (Taylor et al. 2002, 2003). 

These studies assessed effects in adult offspring of rats that were orally exposed during gestation and 

lactation (GD 6 to PND 21). In one study, the maternal rats were dosed with 0, 1, 10, or 30 mg/kg/day 

and offspring were evaluated at various ages (not specified) for end points that included motor activity 

development, auditory startle response, and age at eye opening. There were no alterations in any of these 

indices. The second study also found no changes in motor activity, auditory startle response, or eye 

opening in the offspring following exposure to 1–100 mg/kg/day, although fear conditioning performance 

was altered (Taylor et al. 2003). Other effects of DE-71 in both studies included decreased serum T4

levels in the offspring and dams (see Section 5.2.2.2 Endocrine Effects). 

Delays in reproductive development occurred in male and female Wistar rats that were exposed to the 

pentaBDE commercial mixture DE-71 in studies that are incompletely reported as abstracts (Laws et al. 

2003; Stoker et al. 2003). Male rats were gavaged with 0, 3, 30, or 60 mg/kg/day doses of DE-71 in corn 

oil for 5 days on PNDs 24–28, or for 31 days on PNDs 23–53 (Stoker et al. 2003). Pubertal development 

was delayed as indicated by significant increases in the age at preputial separation at 30 and 60 mg/kg/day 

(2.0 and 2.3 days delay, respectively). Seminal vesical and ventral prostate weights were reduced at 

60 mg/kg/day, although there were no significant changes in testis or epididymal weights or serum 

testosterone levels. Female rats were gavaged with DE-71 in corn oil in doses of 0, 3, 30, or 

60 mg/kg/day for 5 days on PNDs 22–26, or for 20 days on PNDs 22–41 (Laws et al. 2003). Pubertal 

development was delayed as indicated by a small but statistically significant delay in the age of vaginal 

opening (1.8 days delay compared to controls) at 60 mg/kg/day. There were no treatment-related changes 

in weights of female reproductive tissues, and information on levels of female reproductive hormones was 

not reported. As summarized in Endocrine Effects in Section 5.2.2.2, the delays in male and female 

reproductive development were accompanied by changes in serum levels of thyroid hormones (decreased 

T3 and T4, increased TSH). The onset of puberty was also delayed in female offspring (vaginal opening), 

but not in male offspring (preputial separation), of Long Evans rats that were subcutaneously exposed to 

1 or 10 mg/kg/day doses of the congener 2,2’,4,4’,5-pentaBDE (BDE 99) on days 10–18 of gestation 

(Lichtensteiger et al. 2003b). 
PBBs and PBDEs 216 

5. HEALTH EFFECTS 

Testicular and ovarian effects were observed in male and female adult offspring of Wistar rats that were 

maternally exposed to a single 60 or 300 μg/kg oral dose of the pentaBDE congener BDE 99 in peanut oil 

on day 6 of gestation (Kuriyama and Chahoud 2003; Talsness et al. 2003). Evaluations of the male 

offspring were performed on PND 140 and included reproductive tissue weights (testis, epididymis, 

ventral prostate, and seminal vesicle), spermatid and sperm numbers, and serum levels of testosterone and 

LH (Kuriyama and Chahoud 2003). Effects in the male offspring occurred at 60 and 300 μg/kg, including 

significantly decreased (p<0.05) spermatid number (31 and 34% less than controls), sperm number 

(29 and 18% less than controls), and daily sperm production (31 and 34% less than controls). The female 

offspring were necropsied in estrus on approximately postnatal day 90, at which time, ovaries were 

examined by electron microscopy (Talsness et al. 2003). Ultrastructural alterations were observed in 

ovaries at both dose levels, particularly degenerative changes that included vacuolization and 

accumulation of vesicules in the cytoplasm of the stromal cells and necrosis in the serosal epithelial cells. 

The single congener dose levels in this study are much lower than the LOAEL of 2 mg/kg/day for liver 

effects in rats exposed to a commercial mixture of pentaBDE in the 90-day study used to derive the 

intermediate oral MRL. 

The highest NOAEL values and all LOAEL values from each reliable study for developmental effects in 

each species and duration category are recorded in Table 5-3 and plotted in Figure 5-3. 

5.2.2.7 Cancer 

Polybrominated Biphenyls. No epidemiological studies were located that provided evidence for an 

association between exposure to PBBs and the occurrence of cancer in humans, although one case report 

is available concerning gastrointestinal cancer in a Michigan dairy farmer with known exposure to PBBs 

and other chemicals. 

The Michigan Department of Public Health, the U.S. Center for Disease Control, the National Institutes 

of Health, the Food and Drug Administration, and the EPA established a cohort of Michigan residents 

with varying levels of PBB exposure to determine the short- and long-term effects (especially cancer) of 

exposure to PBBs (Landrigan et al. 1979). The epidemiological and clinical data collected during the first 

4 years after the Michigan PBB contamination episode indicated that cancer was not a prevalent 

"symptom" among the cohort at that time. Prevalence rates for cancer in exposed groups ranged from 

0.4 to 0.6% compared with 0% in a small control group comprised of residents of farms with low PBB 

contamination (Landrigan et al. 1979). When the cohort was divided into seven groups based on serum 
PBBs and PBDEs 217 

5. HEALTH EFFECTS 

PBB levels, no trend with concentration was apparent, but the incidence of cancer was the highest in the 

group with the highest serum PBB levels. Subsequent follow-up examinations of this cohort have not 

been reported. 

In studies conducted by the Environmental Science Laboratory of the Mount Sinai School of Medicine, 

mean plasma levels of carcinoembryonic antigen (CEA), which has been used as a screening tool for 

tumor recurrence after cancer therapy, were found to be slightly higher in 1976 in a population of 

611 Michigan residents who likely ingested PBB-contaminated food than mean levels in a nonexposed 

population of Wisconsin farm residents, but the difference was not statistically significant (Anderson et 

al. 1978b). Cancer was not listed as a condition in the report of results of a symptomatology survey 

completed by this cohort (Anderson et al. 1979). Reports of follow-up examinations of this cohort have 

not been reported. 

A relationship between serum PBBs and risk of breast cancer was suggested in a nested case-control 

study of 1,925 women enrolled in the Michigan Department of Public Health registry for persons exposed 

to PBBs (Henderson et al. 1995). Study participants had lived on or received food from a farm 

quarantined by the Michigan Department of Agriculture, were recruited from July 1976 to December 

1977, and followed up annually from 1978 through 1993. Twenty women who developed breast cancer 

were age- and race-matched to 290 controls. Median serum PBB concentrations were similar in the 

cancer cases (3 ppb, range 0.5–16 ppb) and controls (2 ppb, range 0.5–419 ppb). Conditional logistic 

regression was used to calculate odds ratios (ORs) and 95% confidence intervals (CIs) for increasing 

serum PBB levels categorized into tertiles (<2 ppb, 2–3 ppb, ≥4 ppb) and a dichotomous variable (<2 ppb, 

≥2 ppb). The estimated risk for breast cancer was slightly elevated for women with serum PBB levels of 

2–3 ppb (OR=3.5, 95% CI=0.9–13) and ≥4 ppb (OR=3.1, 95% CI=0.8–12) when compared with the 

reference group (<2 ppb), or when the dichotomous variable was used in the analysis (OR=3.3, 95% 

CI=0.9-11.4). The results were essentially the same when the data were adjusted for available risk factors 

(body mass index and family history of cancer), or when matched sets of cases and controls were 

stratified into two groups based on date of diagnosis (≥10 years after exposure). The results of this study 

are inconclusive due to the small number of cases, apparent lack of statistically significant increases 

(p values were not reported, but confidence intervals were broad with lower limits less than unity, 

indicating that it is difficult to exclude chance as an explanation for the findings), and lack of information 

on important breast cancer risk factors (e.g., exposure to other organochlorine chemicals and estrogen 

receptor status). 
PBBs and PBDEs 218 

5. HEALTH EFFECTS 

Another study of the Michigan PBB registry evaluated associations between levels of serum PBBs and 

risks of various site-specific cancers (Hoque et al. 1998). Primary cancers (195 malignancies) were 

identified in 187 persons among 3,899 registrants enrolled in 1976 and followed through 1993. Controls 

were 696 randomly selected cancer-free individuals who were frequency matched to cases by age (in 

5-year strata) and sex in a 4:1 ratio (except above age 70 years when, due to lower numbers, all eligible 

controls were used). PBB levels in the cases were measured at the time of registry enrollment. Serum 

PBB concentration ranges was categorized into four groups (not detectable–3 ppb, 4–20 ppb, 21–50 ppb, 

>50 ppb) defined by the median and the 90th and 95th percentiles. Conditional logistic regression was 

used to calculate univariate and multivariate (adjusted) ORs by cancer site for the three highest serum 

PBB categories compared to the reference (≤3 ppb) group. The multivariate ORs were adjusted for 

family history of cancer, smoking status, alcohol use, age, serum PCB level, and sex. Digestive system 

cancer (12 cases) and lymphoma (not otherwise specified) (8 cases) showed increasing dose-response 

relationships for risk as PBB concentrations increased. Digestive system cancer was a grouping that 

comprised of the following sites: liver (five cases), stomach (five cases), esophagus (one case) and 

pancreas (one case). Adjusted ORs for digestive system cancer were 1.00 (reference), 8.23 (95% 

CI=1.27–53.3), 12.3 (0.80–191), and 22.9 (1.34–392) for the ≤3, 4–20, 21–50, and >50 ppb categories, 

respectively. The corresponding adjusted ORs for PBB level and lymphoma risk were 1.00, 3.85 (0.32–

46.2), 19.6 (1.52–253), and 48.9 (4.09–585). The lymphoma ORs were incompletely adjusted due to 

missing data for serum PCB level and family cancer history in the reference category. Increased risks 

were also observed for breast cancer in the 4–20 ppb category (nine cases, adjusted OR=2.41, 95% 

CI=0.92–6.30), cancer at an unknown site in the 4–20 ppb category (four cases, adjusted OR=31.0, 95% 

CI=1.40–685), and leukemia in the 21–50 ppb category (one case, adjusted OR=4.49, 95% CI=0.92–

6.30). The associations found in this study should be viewed as suggestive and preliminary due to the 

small numbers of cases. The 2.4-fold increased risk of breast cancer for PBB levels between 4 and 

20 ppb is consistent with the 3-fold increased risk for breast cancer observed for PBB levels >2 ppb in the 

Henderson et al. (1995) study of the same cohort summarized above. 

A Michigan dairy farmer, who had a history of health complaints after 1976, developed malignant cancer 

of the esophageal and stomach wall in 1986; the man subsequently died in 1988 (Sherman 1991). 

Samples of adipose tissue, collected in 1976 and 1987, revealed PBB concentrations of 0.83 and 

0.85 ppm, respectively. Also detected in the fat tissue collected in 1987 were polychlorinated biphenyls 

(PCBs) at 3.57 ppm and chlordane residues at concentrations ranging from 0.018 to 0.039 ppm. 
PBBs and PBDEs 219 

5. HEALTH EFFECTS 

FireMaster FF-1 has induced liver tumors and/or neoplastic nodules in rats and mice following single or 

repeated administration by gavage in oil vehicle, as well as following chronic dietary administration. In 

female Sherman rats given a single 1,000 mg/kg dose and observed for 2 years, incidences of 

hepatocellular carcinomas and liver neoplastic nodules were 41.4% (24/58) and 72.4% (42/58), 

respectively (Kimbrough et al. 1981). In an earlier study using the same treatment (single 1,000 mg/kg 

dose), groups of five Sherman rats of each sex were examined at 2, 6, 10, and 14 months following 

treatment (Kimbrough et al. 1978b). Neoplastic nodules were found in the livers of 22.5% (9/40) of the 

treated rats observed for at least 10 months (4/5, 2/5, and 3/5 in the 10-month females, 14-month males, 

and 14-month females, respectively). Liver tumors were not found, but this could be related to the 

relatively small number of animals (20/sex) and/or short duration of observation (≤14 months). Liver 

neoplastic nodules without tumors also developed in 31.2% (5/16) of Sherman rats treated with a lower 

single dose (200 mg/kg) and observed for 18–22 months (Kimbrough et al. 1981). No liver tumors or 

neoplastic nodules developed in untreated control groups in any of these single dose studies, and 

treatment-related tumors in sites other than liver were not observed. When administered to pregnant rats 

once on gestation day 7, a dose of 200 mg/kg induced both hepatocellular carcinomas and liver neoplastic 

nodules in offspring that were observed for 2 years (Groce and Kimbrough 1984). The incidences of 

tumors in male offspring (9.6% [4/41] versus 0% [0/42] in controls) and nodules in female offspring 

(17.6% [9/51) versus 4.2% [2/48]) increased significantly (p≤0.055). Treatment-related tumors were not 

observed in nonhepatic tissues of the offspring. 

Hepatocellular carcinomas and liver neoplastic nodules also increased in female Sherman rats gavaged 

with 100 mg/kg FireMaster FF-1 twice a week for two 3-week periods (12 total doses) separated by 

≈10 weeks (Kimbrough et al. 1981). Following observation for 2 years, the incidence of carcinomas was 

60.7% (17/28 versus 0/25 in controls) and incidence of nodules was 85.7% (24/28 versus 1/25). In 

repeated dose studies performed by NTP (1983), FireMaster FF-1 was administered via gavage to 

Fischer-344/N rats and B6C3F1 mice of both sexes at dosages of 0, 0.1, 0.3, 1, 3, or 10 mg/kg/day on 

5 consecutive days per week for 25 weeks. Both rats and mice were observed for life (up to 23 and 

24 months posttreatment, respectively). Incidences of hepatocellular carcinoma were dose-related and 

significantly (p<0.01) increased in male rats at ≥3 mg/kg/day (0/33 [controls], 2/39, 0/40, 1/33, 7/33, and 

7/31 [high-dose]) and female rats at 10 mg/kg/day (7/20 versus 0/20 in controls). The incidence of 

cholangiocarcinoma was significantly (p<0.01) increased in female rats at 10 mg/kg/day (7/20 versus 

0/20) and almost significant (p=0.06) in males at 10 mg/kg/day (2/31 versus 0/33). Liver neoplastic 

nodules were dose-related and significantly (p<0.01) increased in female rats at ≥3 mg/kg/day. No clear 

treatment-related effects on incidences of hepatic neoplastic nodules in males, or bile duct hyperplasia, 
PBBs and PBDEs 220 

5. HEALTH EFFECTS 

myelomonocytic (mononuclear cell) leukemia, or foci of pancreas-like tissue in the liver in either sex 

were observed. In mice, incidences of hepatocellular carcinoma significantly (p<0.01) increased at 

10 mg/kg/day in males (21/22 versus 12/25 in controls) and females (7/8 versus 0/13). Metastasis to lung 

also significantly (p<0.05) increased in female mice at 10 mg/kg/day. No treatment-related effects on 

hepatocellular adenomas or hepatoblastomas were observed. Thyroid follicular cell adenoma tended to 

increase in treated female mice, but data are inconclusive due to low incidences and small numbers of 

animals. 

The carcinogenicity of FireMaster FF-1 was additionally evaluated in Fischer-344/N rats and B6C3F1 

mice of both sexes that received adult exposure only, perinatal exposure only, or combined perinatal and 

adult exposure (NTP 1992). The adult-only exposure involved dietary administration of PBBs (F1 diets) 

to ≈8-week-old animals for up to 104 weeks (rats) or 105 weeks (mice). Perinatal-only exposure involved 

dietary treatment of dams (F0 diets) for 60 days prior to breeding and throughout gestation and lactation 

until pups were 8 weeks old. The pups were administered the same treatment as the dams from weaning 

at week 4 until age 8 weeks, and were subsequently administered the same or different dietary treatments 

(F1 diets) for up to 104 weeks (rats) or 105 weeks (mice). This study was designed to compare the 

carcinogenicity of PBBs given in a conventional bioassay protocol (i.e., the adult-only exposure) with that 

of PBBs given in a combined perinatal and adult exposure regimen. 

Eight F0:F1 doses (estimated) were tested in rats among one unexposed control group (0:0 mg/kg/day), 

two adult-only exposure groups (0:0.5 and 0:1.5 mg/kg/day), one perinatal-only exposure group 

(0.5:0 mg/kg/day), and four combined perinatal and adult exposure groups (0.05:0.15, 0.15:0.5, 0.5:0.5, 

and 0.5:1.5 mg/kg/day) (Chhabra et al. 1993; NTP 1992). Incidences of hepatocellular tumors were 

increased in adult-only exposed rats of both sexes. In males ingesting 0:0, 0:0.5, and 0:1.5 mg/kg/day, 

incidences of adenoma were 1 of 50, 10 of 49, and 38 of 50; of carcinoma, 0 of 50, 2 of 49, and 19 of 50; 

and of adenoma or carcinoma (combined), 1 of 50, 12 of 49, and 41 of 50, respectively. In females, 

incidences of adenoma were 0 of 50, 10 of 50, and 38 of 50; of carcinoma, 0 of 50, 2 of 50, and 4 of 50; 

and of adenoma or carcinoma (combined), 0 of 50, 12 of 50, and 39 of 50, respectively. These increases 

in liver tumor incidences were statistically significant (p≤0.002) except for carcinoma in 0:0.5 mg/kg/day 

males and females and 0:1.5 mg/kg/day females (p>0.05). Combined perinatal and adult exposure 

significantly enhanced the development of liver tumors in female rats, as shown by comparisons with 

females receiving adult exposure only. Compared to the 0:0.5 mg/kg/day female adult-only exposed 

group, incidences of hepatocellular adenoma were 22 of 50 (p=0.01) and 35 of 50 (p<0.001); of 

carcinoma, 1 of 50 (p=0.05) and 8 of 50 (p=0.048); and of hepatocellular adenoma or carcinoma 
PBBs and PBDEs 221 

5. HEALTH EFFECTS 

(combined), 22 of 50 (p=0.03) and 39 of 50 (p<0.001) in the 0.15:0.5 and 0.5:0.5 mg/kg/day groups, 

respectively. Compared to the 0:1.5 mg/kg/day female adult-only exposed group, incidences of 

hepatocellular adenoma, carcinoma, and hepatocellular adenoma or carcinoma (combined) were 45 of 

50 (p=0.049), 22 of 50 (p<0.001), and 47 of 50 (p=0.016), respectively, in the 0.5:1.5 mg/kg/day group. 

This enhancing influence of perinatal exposure did not occur in the males. Perinatal-only exposure did 

not cause significantly increased incidences of liver or other tumors in rats of either sex. 

Increased incidences of mononuclear cell leukemia occurred in adult-only exposed rats but were not 

clearly related to treatment (Chhabra et al. 1993; NTP 1992). The incidences of this leukemia in the 0:0, 

0:0.5, and 0:1.5 mg/kg/day groups were 25 of 50, 33 of 50, and 31 of 50, respectively, in males and 14 of 

50, 22 of 50, and 23 of 50, respectively, in females; the incidences in the 0:0.5 mg/kg/day males and 

0:1.5 mg/kg/day females were significantly (p≤0.05) increased. Comparison of the combined perinatal 

and adult exposure groups with the adult-only exposed groups showed no significant enhancement; 

however, comparison with the unexposed control (0:0 mg/kg/day) incidences showed a consistent 

increase in the incidence of this neoplasm at higher doses. In the 0.15:0.5 and 0.5:0.5 mg/kg/day groups, 

the incidences of leukemia were 41 of 50 (p≤0.01) and 37 of 50 (p≤0.01) in males and 17 of 50 (p>0.05) 

and 27 of 50 (p≤0.01) in females. In the 0.5:1.5 mg/kg/day groups, the incidences were 37 of 50 (p≤0.01) 

in males and 25 of 50 (p≤0.05) in females. The incidences in some of these groups fall outside the NTP 

historical control range. The incidences in males were as high as 82% and exceeded the upper historical 

control range of 62%. In females, the incidences were as high as 54% and exceeded the overall upper 

historical control range of 52% and the laboratory upper historical control range of 28%. A combined 

(life table) analysis of data from all eight experimental groups indicates that significant increases in the 

incidence of the leukemia are associated with increasing F1 concentrations (p≤0.05 in males; p≤0.01 in 

females). For males, there was also a marginally significant (p≤0.05) increase associated with F0

exposure. 

Eight F0:F1 doses were also tested in the mice among one unexposed control group (0:0 mg/kg/day), two 

adult-only exposure groups (0:1.3 and 0:3.9 mg/kg/day), one perinatal-only exposure group 

(3.9:0 mg/kg/day), and four combined perinatal and adult exposure groups (0.39:0.39, 1.3:1.3, 3.9:1.3, 

and 3.9:3.9 mg/kg/day) (NTP 1992). As in the rats, hepatocellular tumors were significantly (p<0.001) 

increased in adult-only exposed mice of both sexes. In males ingesting 0:0, 0:1.3, and 0:3.9 mg/kg/day, 

incidences of adenoma were 9 of 50, 48 of 49, and 42 of 50; of carcinoma, 8 of 50, 30 of 49, and 36 of 50; 

and of adenoma or carcinoma (combined), 16 of 50, 48 of 49, and 48 of 50, respectively. In adult-only 

exposed females, incidences of adenoma were 4 of 50, 39 of 50, and 46 of 48, carcinoma were 1 of 50, 
PBBs and PBDEs 222 

5. HEALTH EFFECTS 

28 of 50, and 41 of 48, and adenoma or carcinoma (combined) were 5 of 50, 42 of 50, and 47 of 48, 

respectively. Combined perinatal and adult exposure resulted in increased incidences of liver neoplasms 

in some treated groups. However, because adult-only exposure to 1.3 or 3.9 mg/kg/day resulted in such 

high incidences of liver neoplasms (84–98%), the possible enhancing effect of combined perinatal and 

adult exposure could not be adequately assessed in either sex. Compared to 0:1.3 mg/kg/day adult-only 

exposure, combined perinatal and adult exposure to 3.9:1.3 mg/kg/day caused significantly increased 

incidences of carcinoma in males (40 of 50, p=0.01) and females (44 of 50; p<0.001), adenoma in females 

(47 of 50, p=0.005) and adenoma or carcinoma (combined) in females (50 of 50, p<0.001). Compared to 

0:3.9 mg/kg/day adult-only exposure, combined perinatal and adult exposure to 3.9:3.9 mg/kg/day caused 

significantly increased adenoma incidence in males (48 of 50; p=0.007) and decreased adenoma incidence 

in females (41 of 47; p=0.022). Perinatal-only exposure also caused significantly increased incidences of 

liver neoplasms in mice of both sexes. Comparison of the 0:0 and 3.0:0 mg/kg/day groups showed 

hepatocellular adenoma, carcinoma, and adenoma or carcinoma (combined) incidences of 31 of 

50 (p<0.001), 17 of 50 (p=0.033), and 40 of 50 (p<0.001) in males and 19 of 50 (p<0.001), 7 of 

50 (p=0.213), and 21 of 50 (p<0.001) in females. Combined perinatal and adult exposure to 

3.9:1.3 mg/kg/day also caused a significant (p=0.029) increase in the incidence of thyroid follicular cell 

adenoma in male mice (5 of 48) compared to adult-only exposure to 0:1.3 mg/kg/day (0 of 49). This 

incidence of thyroid adenoma exceeds the historical control range of 0–4% in untreated males in NTP 

studies, but the effect was not seen in the higher dose groups (0:3.9 or 3.9:3.9 mg/kg/day). Perinatal-only 

exposure did not induce thyroid or other nonhepatic tumors in mice of either sex. 

The existing evidence conclusively demonstrates that the liver is the main target of PBB carcinogenicity 

in animals. Results of a chronic study (Chhabra et al. 1993; NTP 1992) suggest that male rats are more 

sensitive than female rats (based on a higher carcinoma/adenoma ratio), and that mice are more sensitive 

than rats (based on earlier occurrence of hepatocellular adenomas, higher combined incidence of all liver 

neoplasms, and higher liver concentrations of PBBs). Based on findings in male rats and mice of both 

sexes in this study, there is some evidence that combined perinatal and adult dietary exposure to 

FireMaster FF-1 enhanced the susceptibility of hepatocellular neoplasms in animals receiving adult 

exposure. 

The Cancer Effect Levels (CELs) for FireMaster FF-1 reported in Kimbrough et al. (1981), Groce and 

Kimbrough (1984), and NTP (1983, 1992) are recorded in Table 5-2 and plotted in Figure 5-2. 
PBBs and PBDEs 223 

5. HEALTH EFFECTS 

There are data indicating that FireMaster BP-6 has tumor promoting activity in rats and hamsters. In 

standardized liver tumor promotion assays, development of enzyme-altered hepatic foci (putative 

preneoplastic lesions) was assessed in rats that were 70% hepatectomized, initiated with a 

subcarcinogenic intraperitoneal dose of diethylnitrosamine 24 hours after the partial hepatectomy, and 

promoted with orally administered FireMaster BP-6 beginning 30 days later. Various promotion 

protocols caused significantly increased numbers of enzyme-altered hepatic foci with gamma-glutamyl 

transpeptidase (GGT) activity, including two gavage doses of 65 mg/kg on adjacent days (6.5 mg/kg was 

not effective) (Rezabek et al. 1987), estimated dietary dosages of 0.5 or 5 mg/kg for 180 days (Jensen et 

al. 1982), and estimated dietary dosages of 0.5 mg/kg/day for 140 days or 5 mg/kg for 15 days (Jensen et 

al. 1984). In a similar assay with rats that were not hepatectomized, a single gavage dose of 100 mg/kg 

FireMaster BP-6 administered 7–10 days after initiation with dimethylnitrosamine (NDMA) or N-nitrosopyrrolidine (NPYR) promoted development of hepatic enzyme-altered foci (Rangga-Tabbu and Sleight 

1992). A statistically significant increased number of tracheal papillomas (but not number of animals 

with papillomas) developed in a group of hamsters given a single subcutaneous initiating dose of diethylnitrosamine and fed an estimated dietary dosage of 8.3 mg/kg/day FireMaster BP-6 for 140 days (Wasito 

and Sleight 1989). 

Individual PBB congeners have been examined for tumor promoting activity in rats that were partially 

hepatectomized and initiated with diethylnitrosamine (DENA). Numbers of hepatic GGT-altered foci 

and/or neoplastic nodules were increased following promotion with 3,3',4,4'-tetrabromobiphenyl (BB 77) 

(≈0.25 mg/kg/day in the diet for 180 days or 8 weekly intraperitoneal injections of ≈7 mg/kg), 

2,2',4,4',5,5'-hexabromobiphenyl (BB 153) (≈0.5 mg/kg/day in the diet for 180 days) or 3,3',4,4',5,5'-hexabromobiphenyl (BB 169) (≈0.05 mg/kg/day in the diet for 140 days) (Buchmann et al. 1991; Dixon et al. 

1988; Jensen et al. 1982, 1983). Dietary exposure to ≈5 mg/kg/day of BB 153 for 480 days similarly 

promoted hepatic development of altered foci and neoplastic nodules in rats, whereas ≈0.05 mg/kg/day of 

BB 169 did not, although an apparent synergistic effect was observed when these two congeners were fed 

together (Jensen and Sleight 1986). Additional information on structure-promotion relationships for 

PBBs is discussed in Section 5.5.2. 

The tumor initiating potential of PBBs is not well characterized. Numbers of GGT-altered foci were 

significantly increased in partially hepatectomized rats that were administered a single 1–10 mg/kg oral 

dose of 3,3',4,4'-tetrabromobiphenyl (BB 77) followed by phenobarbital in the diet for 180 days (Dixon et 

al. 1988), indicating that PBBs may have initiating activity in hepatocarcinogenesis. The potential for 
PBBs and PBDEs 224 

5. HEALTH EFFECTS 

liver tumor initiation by PBBs appears to be weak compared to their potent promoting activity 

(Buchmann et al. 1991; Dixon et al. 1988; Jensen et al. 1984). 

Polybrominated Diphenyl Ethers. There was no clear association between risk of non-Hodgkin’s 

lymphoma (NHL) and exposure to 2,2'4,4'-tetraBDE in a case-control study of 77 Swedish men and 

women who were recruited in 1995–1997 and ranged in age from 28 to 85 years (Hardell et al. 1998; 

Lindstrom et al. 1998). Adipose tissue levels of 2,2',4,4'-tetraBDE (BDE 47) (used as a marker for total 

PBDE exposure) were compared in 19 patients with NHL, 23 patients with malignant melanoma, 

8 patients with other cancers or in situ changes, and 27 persons with no cancer diagnosis. The highest 

concentrations were seen in the patients with NHL. The mean concentration of BDE 47 was 13.0 ng/g 

(ppb) lipid (range 1.0–98.2 ppb) in the 19 NHL patients and 5.1 ppb (range 0.6–27.5 ppb) in the 

27 persons without known malignancies. Logistic regression, adjusted for age, gender, sum of PCBs, and 

sum of chlordanes, was performed on cases and controls in three concentration groups (<2.05, 2.05–

<5.43, and ≥5.43 ppb). A nonsignificantly elevated risk with a suggestive dose-response was found for 

NHL in the two highest concentration groups compared with the lowest group; the ORs and 95% CIs 

were 1.9 (0.3–14) and 3.8 (0.7–26) in the middle and high groups, respectively. Although the risk was 

highest in the group with the highest concentration of 2,2'4,4'-tetraBDE (p=0.09 for trend), there was no 

significant difference between cases and controls (p=0.14). The results for patients with malignant 

melanoma did not differ from controls. 

Information on carcinogenic effects of PBDEs in animals is limited to results of chronic bioassays of 

decaBDE mixtures in rats and mice (Kociba et al. 1975; Norris et al. 1975b; NTP 1986). As summarized 

below, these studies provide limited evidence for the carcinogenicity of decaBDE in animals. No 

carcinogenicity studies of octaBDE or pentaBDE were located in the available literature. 

The National Toxicology Program evaluated the carcinogenicity of commercial grade decaBDE (94-97% 

pure, no detected brominated dioxins or furans) in Sprague-Dawley rats (50/sex/dose) and B6C3F1 mice 

(50/sex/dose) that were exposed in the diet for 103 weeks and observed for an additional 0–1 weeks (NTP 

1986). Comprehensive gross and histological examinations were performed on all animals in all dose 

groups including those that were moribund or died during the study. Reported estimated dose levels in 

the rats were 1,120 and 2,240 mg/kg/day in males and 1,200 and 2,550 mg/kg/day in females. Incidences 

of liver neoplastic nodules in low- and high-dose male rats (7/50 and 15/49, respectively) and high-dose 

female rats (9/50) were significantly greater than in controls (1/50 in both males and females) (p≤0.03, 

Fisher Exact test) and showed positive dose-related trends (p<0.001, Cochran-Armitage trend test). 
PBBs and PBDEs 225 

5. HEALTH EFFECTS 

Incidences of hepatocellular carcinoma alone (1/50, control males; 1/50, low-dose males; 1/49, high-dose 

males; 0/50, control females; 2/49, low-dose females; and 0/50, high-dose females) were not significantly 

increased in the treated rat groups compared to controls. The increased incidences of neoplastic nodules 

were considered as “some evidence of carcinogenicity” in both sexes. However, although it was 

concluded that there was some evidence of carcinogenicity in male and female rats based on “neoplastic 

nodules”, this is a poorly defined and understood term that is no longer used by NTP to characterize 

hepatoproliferative lesions in rats. A dose-related trend for mononuclear cell leukemia was observed in 

treated male rats but was not considered to be biologically significant because of a high incidence in 

control animals. 

Reported estimated doses in the mice were 3,200 and 6,650 mg/kg/day in males and 3,760 and 

7,780 mg/kg/day in females (NTP 1986). Hepatocellular adenoma or carcinoma (combined) occurred at 

significantly increased incidences in low-dose male mice (22/50, p=0.002) and high-dose male mice 

(18/50, p=0.019) in comparison to controls (8/50) and showed a positive dose-related trend (p=0.021). 

Incidences of hepatocellular carcinoma alone were not significantly increased in either the low- or highdose male mice. Slightly elevated incidences of thyroid gland follicular cell adenoma or carcinoma 

(combined) were additionally observed in exposed male mice but the increases were not statistically 

significant (control, 0/50; low dose, 4/50; high dose, 3/50). Incidences of follicular cell hyperplasia were 

significantly increased in male mice as summarized in the subsection on Endocrine Effects in 

Section 5.2.2.2. No significantly increased incidences of neoplastic lesions were observed in the female 

mice. NTP (1986) concluded that the significant increase in liver tumors and equivocal increase in 

thyroid tumors represented equivocal evidence of carcinogenicity in male mice. The evidence of 

carcinogenicity in the male mice was considered limited by an early loss of control animals. Losses of 

control male mice were significant during the first year of the study but were subsequently comparable to 

the dosed mice; the early losses were presumed to be due to fighting among animals in both control and 

treatment groups. 

The carcinogenicity of decaBDE was also evaluated in Sprague-Dawley rats (25/sex/dose) that were 

exposed to dietary doses of 0, 0.01, 0.1, or 1.0 mg/kg/day for approximately 2 years (702 days for males, 

735 days for females) (Kociba et al. 1975; Norris et al. 1975b). The commercial mixture was comprised 

of 77.4% decaBDE, 21.8% nonaBDE, and 0.8% octaBDE and therefore differs from typical decaBDE 

formulations containing ≥97% decaBDE. Comprehensive histological examinations showed no 

exposure-related neoplastic effects. The ability of this study to detect carcinogenic changes is limited by 

the very low dose levels in comparison to those tested in the NTP (1986) bioassay. 
PBBs and PBDEs 226 

5. HEALTH EFFECTS 

The Cancer Effect Levels (CELs) for decaBDE in the NTP (1986) study are recorded in Table 5-3 and 

plotted in Figure 5-2. 

5.2.3 Dermal Exposure 

A few studies have examined groups of chemical workers involved in the manufacture and distribution of 

PBBs and/or PBDEs (Bahn et al. 1980; Brown et al. 1981; Chanda et al. 1982; Landrigan et al. 1979; 

Rosenman et al. 1979; Stross et al. 1981). Although the route of exposure (inhalation relative to dermal) 

of these workers has not been well defined, they appear to have had a high potential for dermal exposure 

(Anderson et al. 1978d). Results from these studies are discussed in this section, as well as in 

Section 5.2.1. Dermal exposure may not be an important route of concern for PBDEs because dermal 

absorption is likely to be low, particularly for the highly brominated congeners. 

5.2.3.1 Death 

Polybrominated Biphenyls. No reports of death in humans after dermal exposure to PBBs were located 

in the available literature. 

No deaths were observed over a 14-day period among a group of four rabbits exposed to up to 

10,000 mg/kg of body weight of a commercial octabromobiphenyl mixture by application to abraded and 

occluded dorsal trunk skin (Waritz et al. 1977). The bromobiphenyl was applied as a 35% (w/v) paste in 

corn oil. The same group of investigators reported that four of four rabbits died over a 14-day period after 

application of 5,000 mg/kg of a commercial hexabromobiphenyl mixture in the same vehicle as the 

octabromobiphenyl mixture. A dose of 10,000 mg/kg applied for 24 hours killed two of four rabbits. The 

cause of death was not reported. A commercial mixture of decabromobiphenyl in corn oil was not lethal 

in rats that were observed for 14 days following application of a single dose as high as 5,000 mg/kg to 

covered intact skin (Millischer et al. 1980). The octabromobiphenyl LOAEL of 5,000 mg/kg is reported 

in Table 5-5. It is unclear whether the different lethality rates observed among the hexa-, octa-, and 

decabromobiphenyl mixtures reflect differences in lethal potency or in absorption rates or both. 

Polybrominated Diphenyl Ethers. No reports of death in humans after dermal exposure to PBDEs were 

located in the available literature. 
LOAEL

Table 5-5 Levels of Significant Exposure to Polybrominated Biphenyls - Dermal

Species

(Strain) System

NOAEL Less Serious Serious 

Exposure/

Duration/

Frequency

(Specific Route)

ACUTE EXPOSURE

Death

(NZW)Rabbit 10000

mg/kg/day

mg/kg/day

W

2120 (O

24 hr

(NZW)Rabbit

(4/4 died)

5000

mg/kg/day

mg/kg/day

W

233 (H

24 hr

Systemic

(Albino)Rabbit Dermal 658

mg/kg/day

mg/kg/day

M

389 (D

once

(NZW)Rabbit Dermal

(hyperkeratosis)

0.19

mg/kg/day

mg/kg/day

Ne

272 (F

5 d1x/d

(NZW)Rabbit Hepatic 100 M

mg/kg/day

mg/kg/day

(increased liver weight;necrotic foci)

1000

mg/kg/day

mg/kg/day

W

234 (H

24 hr

Bd Wt 1000 M

mg/kg/day

mg/kg/day

(11% weight loss)

5000

mg/kg/day

mg/kg/day

(NS)Rabbit Hepatic

(increased liver weight)

1

mg/kg/day

mg/kg/day

W(O

240

2 wk5 d/wk1x/d

Dermal 1

mg/kg/day

mg/kg/day

Bd Wt 1

mg/kg/day

mg/kg/day
LOAEL

Table 5-5 Levels of Significant Exposure to Polybrominated Biphenyls - Dermal (continued)

Species

(Strain) System

NOAEL Less Serious Serious 

Exposure/

Duration/

Frequency

(Specific Route)

INTERMEDIATE EXPOSURE

Systemic

(AlbinoHartley)Gn Pig Dermal 62 M

mg/kg/day

mg/kg/day

Bd Wt = body weight; d = days; DBB = deca-brominated biphenyl; Derm = dermal; FF-1 = FireMaster FF-1; HBB = hexa-brominated biphenyl; hr = hour(s); M 

octa-brominated biphenyl; wk = week(s); x = time(s)

W(O

238

3 wk3x/wk
PBBs and PBDEs 229 

5. HEALTH EFFECTS 

No deaths occurred in rabbits that were observed for 14 days following a single ≤2,000 mg/kg dermal 

dose of decaBDE, octaBDE or pentaBDE (IRDC 1974, 1975a, 1975b). The PBDEs were applied to 

clipped intact skin, covered with an occlusive barrier, and washed from the treatment site 24 hours later. 

5.2.3.2 Systemic Effects 

No studies were located regarding respiratory, cardiovascular, gastrointestinal, hematological, 

musculoskeletal, or renal effects in humans or animals after dermal exposure to PBBs. 

No studies were located regarding respiratory, cardiovascular, gastrointestinal, hematological, 

musculoskeletal, hepatic, or renal effects in humans or animals after dermal exposure to PBDEs. 

Systemic effects that have been observed in humans and animals following dermal exposure to PBBs and 

PBDEs are described below. The highest NOAEL and all LOAEL values from each reliable study for 

systemic end points in each species and duration category are recorded in Tables 5-5 and 5-6. 

Hepatic Effects. 

Polybrominated Biphenyls. No studies were located regarding hepatic effects in humans after dermal 

exposure to PBBs. 

No significant changes in relative or absolute liver weight or gross pathological effects were reported in 

groups of four rabbits after application of a single dose of up to 10,000 mg/kg of octabromobiphenyl 

mixture in corn oil to abraded and occluded dorsal skin over a 24-hour period (Waritz et al. 1977). It was 

unclear if histopathological examinations were performed. Using the same protocol in rabbits, these 

investigators reported a significant increase (p<0.01) in relative and absolute liver weight, distinct lobular 

markings, and necrotic foci with doses ≥1,000 mg/kg of a commercial hexachlorobiphenyl mixture. A 

dose of 100 mg/kg was without effect. A significant increase (p<0.01) in relative liver weight was 

reported in rabbits after application of 1 mg/kg/day of a commercial mixture of octabromobiphenyl in 

corn oil to the intact and occluded shaved dorsal skin on 5 days/week for 2 weeks (Waritz et al. 1977). 

Histopathological examinations were not performed. A relatively low dose (0.0013 mg/kg) of FireMaster 

BP-6 dissolved in benzene/decame (1:9) applied once a day for 5 days to the ear of three rabbits caused 

no histopathological effects in the liver (Hass et al. 1978). 
LOAEL

Table 5-6 Levels of Significant Exposure to Polybrominated Diphenyl Ethers - Dermal

Species

(Strain) System

Reference

NOAEL Less Serious Serious Chemical Form

Exposure/

Duration/

Frequency

(Specific Route)

ACUTE EXPOSURE

Systemic

(New Zealand)

Rabbit Bd Wt 2000

mg/kg

mg/kg

IRDC 1974

2101 DecaBDE

24 hr

(New

Zealand)

Rabbit Bd Wt 2000

mg/kg

mg/kg

IRDC 1975a

OctaBDE

2103

24 hr

(New

Zealand)

Rabbit Bd Wt 2000

mg/kg

mg/kg

Bd Wt = body weight; d = days; Derm = dermal; hr = hour(s); M = male; wk = week(s); x = time(s).

IRDC 1975b

PentaBDE

2105

24 hr

PBBs and PBDEs

5. HEALTH EFFECTS

230
PBBs and PBDEs 231 

5. HEALTH EFFECTS 

No studies were located regarding hepatic effects following intermediate or chronic dermal exposure to 

PBBs. 

Endocrine Effects. 

Polybrominated Biphenyls. Hypothyroidism was diagnosed in 4 of 35 men who were occupationally 

exposed to unspecified PBBs and/or decaBDE (Bahn et al. 1980). The cohort consisted of workers (mean 

age 35.9 years) who had been employed at a production plant for at least 6 weeks during a 52-month 

period during which PBBs and decaBDE were the only chemicals manufactured and who had volunteered 

for a comprehensive medical evaluation performed 3 months after the end of the 52-month period. There 

was no further description of exposure, and it was assumed to have occurred by both inhalation and 

dermal routes. As detailed in Section 5.2.1.2, the results of this study suggest that occupational exposure 

to PBBs, decaBDE, and/or bromine affected the thyroid, but the mixed chemical exposure and a lack of 

data on serum or tissue levels of the chemicals preclude attributing effects solely to any particular 

congener or mixture of congeners. 

No studies were located regarding endocrine effects in animals after dermal exposure to PBBs. 

Polybrominated Diphenyl Ethers. There is suggestive evidence of hypothyroidism in a small group of 

workers who were occupationally exposed to decaBDE as well as PBBs (Bahn et al. 1980) as summarized 

above and detailed in Section 5.2.1.2. 

No studies were located regarding endocrine effects in animals after dermal exposure to PBDEs. 

Dermal Effects. 

Polybrominated Biphenyls. As discussed in Section 5.2.1.2, results from a medical history survey study 

of workers in a PBB manufacturing plant and a nonexposed group of Wisconsin farm residents indicated 

an association between occupational exposure to PBBs and the occurrence of acne (Chanda et al. 1982). 

The survey covered a period of 3 years of potential exposure, but exposure levels were not reported. No 

adverse dermal effects were observed on the arms or legs of subjects after a 6-day application of polymer 

fibers containing commercial octabromobiphenyl mixture under an occlusive covering; no additional 

information was reported (Waritz et al. 1977). 
PBBs and PBDEs 232 

5. HEALTH EFFECTS 

Several studies examined the acute dermal effects of commercial PBB mixtures in rabbits. Application of 

0.19 mg/kg FireMaster FF-1 for 5 days, in toluene vehicle, to the inner surface of the left ear of two 

rabbits (right ear served as control) induced moderate hyperkeratosis, which included marked dilation of 

the hair follicles, with moderate proliferation of the epithelium and partial atrophy of the sebaceous 

glands (Needham et al. 1982). There was also evidence of excess keratin and debris in the subjacent hair 

follicles. Application of either a dry or water-moistened formulation of octabromobiphenyl mixture 

(amount not reported) for 24 hours did not adversely affect intact skin in rabbits, but slight erythema and 

edema were observed in abraded skin (Norris et al. 1975a). Repeated applications over a 2-week period 

of the dry octabromobiphenyl mixture formulation (amount not reported) to occluded intact or abraded 

skin caused no skin response, but the water-moistened formulation caused slight and transient erythema 

(Norris et al. 1975a). None of these studies reported the number of animals used. Rough skin with mild 

erythema was observed in occluded intact shaved dorsal skin of a group of rabbits after repeated 

applications of a dose of 1 mg/kg/day octabromobiphenyl mixture in corn oil over a 2-week period 

(Waritz et al. 1977). Application of a commercial decabromobiphenyl mixture in olive oil to covered 

intact or abraded skin for 4 hours, in an amount equivalent to 658 mg/kg, caused very slight erythema 

with or without edema in rabbits (Millischer et al. 1980). 

Limited information is available regarding intermediate-duration dermal effects of PBBs in animals. A 

10% chloroform solution of an unspecified commercial formulation of octabromobiphenyl did not induce 

bromacne when applied to the ear of rabbits for 30 days (Norris et al. 1975a). Only slight erythema and 

exfoliation was observed. Doses of 62 mg/kg of octabromobiphenyl mixture were not sensitizing when 

applied to the intact or abraded skin of guinea pigs over a 3-week period (Waritz et al. 1977). 

No studies were located regarding dermal effects in animals after chronic application of PBBs. 

Polybrominated Diphenyl Ethers. A human sensitization study was conducted in which 77.4% decaBDE 

(containing 21.8% nonaBDE and 0.8% octaBDE ) as a 5% suspension in petrolatum was applied via 

patch, 3 times a week for 3 weeks, to 50 subjects (Norris et al. 1975a). No skin sensitization responses 

occurred during the sensitizing period or on challenge 2 weeks following the last application. No 

additional information was reported regarding the design and results of this study. 

There was no evidence of primary irritation in intact skin of rabbits that were dermally exposed to a 

former commercial decaBDE mixture (500 mg as dry solid was applied to clipped skin and occluded for 

24 hours) (IRCD 1974). A similar application of 77.4% decaBDE (containing 21.8% nonaBDE and 0.8% 
PBBs and PBDEs 233 

5. HEALTH EFFECTS 

octaBDE) (dry solid, amount not reported), octaBDE (500 mg as dry solid), or pentaBDE (0.5 mL as a 

viscous liquid) was also non-irritating to intact rabbit skin (IRCD 1975a, 1975b; Norris et al. 1975a). A 

similar single application of solid 77.4% decaBDE (21.8% nonaBDE and 0.8% octaBDE) to abraded skin 

caused a slight erythematous and edematous response in rabbits, although repeated applications of the 

same decaBDE mixture to intact skin (5 days/week for 2 weeks) or abraded skin (3 days) was nonirritating to rabbits (Norris et al. 1975a). All skin sites returned to normal appearance following cessation 

of exposure. (Norris et al. 1975a). 

OctaBDE and pentaBDE were non-sensitizing in maximization tests in guinea pigs (Microbiological 

Associates Inc. 1996). The induction doses consisted of three pairs of interscapular region intradermal 

injections of (1) a 50:50 solution of Freund’s adjuvant and corn oil, (2) 2.5% octaBDE or 5% pentaBDE 

solutions in corn oil, and (3) 2.5% octaBDE or 5% pentaBDE in the 50:50 corn oil/Freund’s adjuvant 

solution. Control groups received the same regimen without PBDEs. After 7 days, the PBDE-treated 

animals received topical applications of neat octaBDE or pentaBDE on the previously treated 

interscapular sites. Two weeks later, the animals were challenged with topical doses of neat octaBDE or 

pentaBDE on the left flank. Subsequent examination of the test sites at 24, 48, 72, 96, or 120 hours after 

the challenge dose showed no erythema or edema responses in any of the animals, indicating that the 

PBDEs did not cause delayed contact hypersensitivity. 

A 10% chloroform solution of 77.4% decaBDE (containing 21.8% nonaBDE and 0.8% octaBDE) did not 

induce bromacne when applied to the ear of rabbits for 30 days (Norris et al. 1975b). A slight 

erythematous response and slight exfoliation were the only observed effects. No additional information 

was reported on the design and results of this acnegenesis study. 

Ocular Effects. 

Polybrominated Biphenyls. No studies were located regarding ocular effects in animals after dermal 

exposure to PBBs. 

Transient irritation of the conjunctival membranes was observed after a single instillation of an 

unreported amount of dry solid octabromobiphenyl mixture to the eye in rabbits, but the cornea, iris, and 

lens were unaffected (Norris et al. 1975a). Commercial grade decabromobiphenyl did not cause eye 

irritation in rabbits when 0.05 mg in olive oil was instilled for 30 seconds followed by rinsing, but 

application of an unspecified amount of dry powder without rinsing was slightly irritating (Millischer et 

al. 1980). Mild conjunctival redness and swelling and a copious discharge was reported after application 
PBBs and PBDEs 234 

5. HEALTH EFFECTS 

of 100 mg of an unspecified commercial PBB powder mixture (either hexa- or octabromobiphenyl) for 

20 seconds into the conjunctival sac of two rabbits (Waritz et al. 1977). These effects disappeared within 

4 hours in both washed (with tap water) and unwashed eyes. The iris and cornea were unaffected. 

Polybrominated Diphenyl Ethers. No studies were located regarding ocular effects in animals after 

dermal exposure to PBDEs. 

Ocular effects were investigated in rats that had 100 mg decaBDE (solid), 100 mg octaBDE (solid), or 

0.1 mL pentaBDE (viscous liquid) instilled into the conjunctival sac (ICRD 1974, 1975a, 1975b). The 

eyes were examined for irritation after 24, 48, and 72 hours and 7 days and corneal injury after 72 hours. 

There were no exposure-related effects with decaBDE or octaBDE, although pentaBDE caused slight 

evidence of corneal damage in one of six rats (IRDC 1975b). 

Body Weight Effects. 

Polybrominated Biphenyls. No studies were located regarding body weight effects in humans after 

dermal exposure to PBBs. 

No treatment-related effects on body weight were reported in rabbits given a dose of 1 mg/kg/day of a 

commercial mixture of octabromobiphenyl in corn oil via application to the intact shaved dorsal skin for 

2 weeks (Waritz et al. 1977). No significant effect on final body weight was reported in groups of four 

rabbits after application of a single dose of up to 10,000 mg/kg of octabromobiphenyl mixture in corn oil 

to the abraded and occluded dorsal skin over a 24-hour period (Waritz et al. 1977). The observation 

period was 14 days. In a similar study with a commercial mixture of hexabromobiphenyl, rabbits treated 

with 1,000 mg/kg showed no weight gain over 14 days. Doses of 5,000 and 10,000 mg/kg induced an 

11% and 20% weight loss, respectively, whereas, a dose of 100 mg/kg was without effect (Waritz et al. 

1977). 

Polybrominated Diphenyl Ethers. No studies were located regarding body weight effects in humans 

after dermal exposure to PBDEs. 

There were no adverse effects on body weight in rabbits that were observed for 14 days following a single 

≤2,000 mg/kg dermal dose of decaBDE, octaBDE, or pentaBDE (IRDC 1974, 1975a, 1975b). The 

PBDEs were applied to clipped intact skin, covered with an occlusive barrier, and washed from the 

treatment site after 24 hours. 
PBBs and PBDEs 235 

5. HEALTH EFFECTS 

5.2.3.3 Immunological and Lymphoreticular Effects 

Polybrominated Biphenyls. Several immunological parameters were evaluated in a group of 28 workers 

from the Michigan Chemical Company who were involved in the manufacturing and distribution of PBBs 

including FireMaster FF-1 (Stross et al. 1981). It is assumed that the main route of exposure was dermal, 

but inhalation and/or oral exposure cannot be ruled out. The subjects had worked directly with PBBs 

during the previous 5 years. Immunological studies included determination of immunoglobulin levels, 

skin testing, and lymphocyte transformation studies. No abnormalities in lymphocyte number or function 

could be determined when compared to an unexposed group. One of three blastogenic responses (PWM) 

was significantly reduced (p<0.01) relative to controls, but was within the normal range for the 

laboratory. No specific information was provided regarding the skin testing and immunoglobulin levels. 

No studies were located regarding immunological effects in animals after dermal exposure to PBBs. 

Polybrominated Diphenyl Ethers. No studies were located regarding immunological effects in humans 

or animals after dermal exposure to PBDEs. 

5.2.3.4 Neurological Effects 

Polybrominated Biphenyls. Twenty-five workers at a PBB manufacturing plant displayed mean scores 

on tests of memory and learning that were typical for people of their age, educational, occupational, and 

cultural backgrounds, even though they displayed an elevated mean PBB concentration in adipose tissue 

(9.33 ppm compared with 3.94 ppm for farm residents) (Brown et al. 1981). Workers with the highest 

concentrations of PBB in adipose tissue showed no evidence of memory dysfunction in these tests. 

Because 15/25 "directly handled PBBs or performed maintenance work in the area where PBBs were 

manufactured," it is likely that at least part of the exposure was by the dermal route. 

No studies were located regarding neurological effects in animals after dermal exposure to PBBs. 

Polybrominated Diphenyl Ethers. No studies were located regarding neurological effects in humans or 

animals after dermal exposure to PBDEs. 
PBBs and PBDEs 236 

5. HEALTH EFFECTS 

5.2.3.5 Reproductive Effects 

Polybrominated Biphenyls. Eleven workers in a PBB manufacturing company in Michigan displayed no 

differences in the distribution of sperm counts, motility, or morphology compared with a control group of 

52 unexposed men (Rosenman et al. 1979). PBBs were detected in the serum of all exposed subjects and 

in only one unexposed subject, but mean or individual serum PBB values were not reported. 

No studies were located regarding reproductive effects in animals after dermal exposure to PBBs. 

Polybrominated Diphenyl Ethers. No studies were located regarding reproductive effects in humans or 

animals after dermal exposure to PBDEs. 

5.2.3.6 Developmental Effects 

Polybrominated Biphenyls. No studies were located regarding developmental effects in humans or 

animals after dermal exposure to PBBs. 

Polybrominated Diphenyl Ethers. No studies were located regarding developmental effects in humans or 

animals after dermal exposure to PBDEs. 

5.2.3.7 Cancer 

Polybrominated Biphenyls. No studies were located regarding cancer in humans after dermal exposure 

to PBBs. 

An unspecified PBB mixture (purity not reported) was not tumorigenic when applied to the shaved dorsal 

skin of female CD-1 mice at a dose of 3.3 mg/kg twice weekly for 30 weeks; no tissues other than skin 

were examined (Berry et al. 1978, 1979). This same treatment did not promote the development of skin 

tumors in mice pretreated with a single application of a tumor initiator, dimethylbenzanthracene 

(DMBA), 1 week prior to PBB exposure (Berry et al. 1978, 1979). The results of these studies must be 

interpreted with caution, since a dose-response study was not done (i.e., only one dose level was tested, 

and the doses may have been too low). Toxic doses of FireMaster FF-1 promoted development of skin 

tumors in female HRS/J hairless mice (Poland et al. 1982). A single dermal application of N-methyl-N'-

nitro-N-nitrosoguanidine (MNNG) initiator, followed by twice weekly applications of FireMaster FF-1 at 
PBBs and PBDEs 237 

5. HEALTH EFFECTS 

66.7 mg/kg for 5 weeks and then 33.3 mg/kg for 15 weeks, resulted in a 60% (9/15) incidence of 

papillomas compared to 0% (0/23) in MNNG-only controls. Toxic effects included mortality, which 

caused the dose reduction after 5 weeks, and severe hepatomegaly and hepatic porphyria. 

Polybrominated Diphenyl Ethers. No studies were located regarding cancer in humans or animals after 

dermal exposure to PBDEs. 

5.3 GENOTOXICITY 

Polybrominated Biphenyls. No studies were located regarding genotoxic effects in humans following 

inhalation, oral, or dermal exposure to PBBs. 

In vivo genotoxicity studies of PBBs in animals are summarized in Table 5-7. Administration of single 

oral doses between 50 and 1,000 mg of FireMaster FF-1/kg (purity not reported) by gavage in corn oil to 

male and female B6C3F1 mice and male Fischer-344 rats did not induce unscheduled deoxyribonucleic 

acid (DNA) synthesis in hepatocytes (Mirsalis et al. 1985, 1989). However, doses ≥200 mg/kg 

significantly increased hepatic cell proliferation in mice, but not in rats. The increase in cell proliferation 

without a change in unscheduled DNA synthesis suggests that PBBs acted as a promoter rather than 

directly causing DNA damage (initiator). A commercial mixture of decabromobiphenyl did not induce 

gene mutation in Salmonella typhimurium bacteria that were intraperitoneally injected into male CFLP 

mice in a host-mediated assay (Millischer et al. 1980). This decabromobiphenyl mixture also did not 

induce micronuclei in bone marrow erythrocytes of mice (Millischer et al. 1980). The mice in the hostmediated assay and micronucleus test were orally treated (method not specified) with total doses of 5,000, 

10,000, or 20,000 mg/kg, administered in two equal doses 24 hours apart. 

In vitro studies indicate that PBBs are not directly genotoxic. As summarized in Table 5-8, PBBs did not 

exhibit mutagenic activity when tested in the prokaryotic organisms S. typhimurium (Haworth et al. 1983; 

Millischer et al. 1980; NTP 1983) and E. coli (Rossman et al. 1991) with or without activation systems in 

the limited number of studies available. In vitro testing in eukaryotic cells resulted in negative genotoxic 

responses in hamster cells (Galloway et al. 1987; Kavanagh et al. 1985; Williams et al. 1984), rat liver 

cells (Kavanagh et al. 1985; Williams et al. 1984), mouse liver and lymphoma cells (Myhr and Caspary 

1991; Williams et al. 1984), and human fibroblasts (Williams et al. 1984). 
PBBs and PBDEs 238 

5. HEALTH EFFECTS 

Table 5-7. Genotoxicity of PBBs and PBDEs In Vivo

Species (test system) End point Results Reference 

PBBs 

Mammalian cells 

 Rat hepatocytes Unscheduled DNA 

synthesis 

– Mirsalis et al. 1989 (FF-1) 

 Mouse hepatocytes Unscheduled DNA 

synthesis 

– Mirsalis et al. 1985, 1989 (FF-1) 

Host-mediated assays: 

Salmonella typhimurium (mouse 

hosted-mediated) 

Gene mutation – Millischer et al. 1980 (DBB) 

Micronucleus test: 

 Mouse bone marrow 

erythrocytes 

Chromosome aberration 

(micronuclei) 

– Millischer et al. 1980 (DBB) 

PBDEs 

Cytogenicity 

 Rat bone marrow cells (onegeneration reproduction study) Chromosome aberration – Norris et al. 1973, 1975 (77.4% 

decaBDE, 21.8% nonaBDE) 

– = negative result; DBB = decabromobiphenyl commercial mixture; decaBDE = decabromodiphenyl ether; 

DNA = deoxyribonucleic acid; FF-1 = FireMaster FF-1; nonaBDE = nonabromodiphenyl ether; 

PBBs = polybrominated biphenyls; PBDEs = polybrominated diphenyl ethers 
PBBs and PBDEs 239 

5. HEALTH EFFECTS 

Table 5-8. Genotoxicity of PBBs and PBDEs In Vitro

Results 

Species (test system) End point 

With 

activation 

Without 

activation Reference 

PBBs 

Prokaryotic organisms: 

Salmonella typhimurium

(plate incorporation) 

Gene mutation – – NTP 1983 (FF-1) 

Salmonella typhimurium

(plate incorporation) 

Gene mutation – – Haworth et al. 1983 (HBB) 

Salmonella typhimurium

(plate incorporation) 

Gene mutation – – Millischer et al. 1980 (DBB) 

Escherichia coli (culture) Gene mutation – – Rossman et al. 1991 (PBB) 

Eukaryotic organisms 

Mammalian cells 

 Chinese hamster CHO 

cells (cell culture) 

Chromosomal 

aberration 

– – Galloway et al. 1987 (HBB) 

 Chinese hamster CHO 

cells (cell culture) 

Sister chromatid 

exchange 

– – Galloway et al. 1987 (HBB) 

 Chinese hamster V79 

cells (cell culture) 

Gene mutation – – Kavanagh et al. 1985 (BP-6) 

 Chinese hamster V79 

cells (cell culture) 

Gene mutation No data – Kavanagh et al. 1985 

(2,4,5-HBB) 

 Chinese hamster V79 

cells (cell culture) 

Gene mutation No data – Kavanagh et al. 1985 

(3,4,5-HBB) 

 Chinese hamster V79 

cells (cell culture) 

Gene mutation – – Kavanagh et al. 1985 

(3,4-TBB) 

 Rat liver cells WB (cell 

culture) 

Gene mutation No data – Kavanagh et al. 1985 

(2,4,5-HBB) 

 Rat liver cells WB (cell 

culture) 

Gene mutation No data – Kavanagh et al. 1985 

(3,4,5-HBB) 

 Mouse lymphoma cells 

L5178Y (cell culture) 

Gene mutation – – Myhr and Caspary 1991 

(FF-1) 

 Rat liver cells (cell 

culture) 

DNA repair No data – Williams et al. 1984 (FF-1) 

 Mouse liver cells (cell 

culture) 

DNA repair No data – Williams et al. 1984 (FF-1) 

 Hamster liver cells (cell 

culture) 

DNA repair No data – Williams et al. 1984 (FF-1) 

 Rat liver cells (cell 

culture) 

Gene mutation No data – Williams et al. 1984 (FF-1) 

 Human fibroblast D-550 

(cell culture) 

Gene mutation – No data Williams et al. 1984 (FF-1) 
PBBs and PBDEs 240 

5. HEALTH EFFECTS 

Table 5-8. Genotoxicity of PBBs and PBDEs In Vitro

Results 

Species (test system) End point 

With 

activation 

Without 

activation Reference 

PBDEs 

Prokaryotic organisms: 

Salmonella typhimurium

(plate incorporation) 

Gene mutation – – NTP 1986 (decaBDE) 

Mammalian cells: 

 Mouse lymphoma 

L5178Y cells (cell 

culture) 

Gene mutation – – NTP 1986 (decaBDE) 

 Chinese hamster 

Sp5/V79cells (cell 

culture) 

Gene recombination No data – Helleday et al. 1999 

(2,2',4,4'-tetraBDE) 

 Chinese hamster 

SPD8/V79cells (cell 

culture) 

Gene recombination No data + Helleday et al. 1999 

(2,2',4,4'-tetraBDE) 

 Chinese hamster 

Sp5/V79 cells (cell 

culture) 

Gene recombination No data + Helleday et al. 1999 

(3,4-diBDE) 

 Chinese hamster 

SPD8/V79 cells (cell 

culture) 

Gene recombination No data + Helleday et al. 1999 

(3,4-diBDE) 

 Chinese hamster 

Sp5/V79 cells (cell 

culture) 

Gene recombination No data + Helleday et al. 1999 

(2-monoBDE) 

 Chinese hamster 

SPD8/V79 cells (cell 

culture) 

Gene recombination No data + Helleday et al. 1999 

(2-monoBDE) 

 Chinese hamster ovary 

cells (cell culture) 

Sister chromatid 

exchange 

– – NTP 1986 (decaBDE) 

 Chinese hamster ovary 

cells (cell culture) 

Chromosomal 

aberrations 

– – NTP 1986 (decaBDE) 

– = negative result; 2,2',4,4'-tetraBDE = 2,2',4,4'-tetrabromodiphenyl ether; 2,4,5-HBB = 2,2',4,4',5,5'-hexabromobiphenyl; 2-monoBDE = 2-bromodiphenyl ether; 3,4,5-HBB = 3,3',4,4',5,5'-hexabromobiphenyl; 3,4-diBDE = 3,4-dibromodiphenyl ether; 3,4-TBB = 3,3',4,4'-tetrabromobiphenyl; BDE = brominated diphenyl ethers; BP-6 = FireMaster 

BP-6; DBB = decabromobiphenyl commercial mixture; decaBDE = decabromobiphenyl ether; DNA = deoxyribonucleic acid; FF-1 = FireMaster FF-1; HBB = hexabromobiphenyl (unspecified); PBB = unspecified mixture; 

PBBs = polybrominated biphenyls; PBDEs = polybrominatedd diphenyl ethers 
PBBs and PBDEs 241 

5. HEALTH EFFECTS 

An in vitro study with a 14C-PBB mixture containing 12 major components found only traces of 

radioactivity bound to rat liver microsomal macromolecules (Dannan et al. 1978a). Binding, however, 

was dependent upon the type of microsomes used to activate the PBB mixture. Microsomes isolated from 

animals pretreated with methylcholanthrene (MC) bound twice the amount of radioactivity compared with 

controls, whereas activation with phenobarbital (PB) or PBB bound 5 times more radioactivity than 

control microsomes. Also, the authors showed that no radioactivity was covalently bound to DNA 

following incubation with 14C-PBB. The type of microsomes used or the presence or bsence of 

nicotinamide adenine dinucleotide phosphate (reduced form) (NADPH) in the incubation mixture made 

no difference. 

Although it appears that PBBs are not mutagenic, due to their enzyme induction properties, they may 

potentiate the genotoxic activity of other compounds by activation to reactive intermediates. 

Polybrominated Diphenyl Ethers. No studies were located regarding genotoxic effects in humans 

following inhalation, oral, or dermal exposure to PBDEs. 

A limited amount of information has been published on the genotoxicity of PBDEs in animals in vivo or 

in prokaryotic and eukaryotic cells in vitro as summarized in Tables 5-7 and 5-8, respectively. 

Cytogenetic examination of bone marrow cells showed no increase in aberrations in maternal and 

neonatal rats following maternal oral exposure to ≤100 mg/kg/day of a 77.4% decaBDE mixture 

(containing 21.8% nonaBDE and 0.8% octaBDE) for 90 days prior to mating and during mating, 

gestation, and lactation (Norris et al. 1973, 1975a). In vitro assays found that decaBDE did not induce 

gene mutations in bacterial cells (S. typhimurium TA98, TA100, TA1535, or TA1537) or mammalian 

cells (mouse lymphoma L5178Y cells), and did not induce sister chromatid exchange or chromosomal 

aberrations in Chinese hamster ovary cells (NTP 1986). In vitro exposure to the congeners 

2,2',4,4'-tetraBDE (BDE 47), 3,4-diBDE (BDE 12), and 2-monoBDE (BDE 1) caused increased 

recombinogenic activity at the HGPRT locus in Chinese hamster SPD8 and Sp5 V79 cells (Helleday et al. 

1999). 

5.4 TOXICOKINETICS 

Data regarding the toxicokinetics of PBBs in humans are limited to information derived from cases of 

accidental ingestion of food contaminated with PBBs and cases of occupational exposure by the 
PBBs and PBDEs 242 

5. HEALTH EFFECTS 

inhalation and dermal routes. These data provide qualitative evidence that PBBs are absorbed in humans 

by the inhalation, oral, and dermal routes. Limited quantitative data in animals indicate that some PBB 

congeners are well absorbed after oral exposure. Dermal absorption data for animals are insufficient for 

estimating absorption rates, and no inhalation absorption data were located. In blood, 80% of PBBs, are 

bound to protein and 20% are associated with lipids. The distribution pattern of PBBs did not differ 

significantly between humans and animals and among animal species. Due to their lipophilic nature, 

PBBs, especially the highly brominated congeners, tend to accumulate in lipid-rich tissues. Greater 

relative amounts of PBBs are usually found in the liver, adipose, skin, and breast milk. Certain 

components of PBB mixtures are metabolized by the microsomal monooxygenase system catalyzed by 

cytochrome P-450 of the type induced by phenobarbital. The rate of metabolism of some PBB congeners 

depends on the bromine substitution pattern. PBB congeners of low bromine content are transformed into 

hydroxylated derivatives that are predominately eliminated in the urine. Highly brominated congeners are 

either retained or excreted unchanged in the feces. As discussed in Section 5.3.2, the exact mechanism of 

PBB toxicity is not known. It has been suggested, however, that the mechanism for some congeners is 

related to the enhancement of gene expression triggered by initial binding to the same cytosolic receptor 

(Ah) involved in some effects of PCBs and PCDDs. 

5.4.1 Absorption 

5.4.1.1 Inhalation Exposure 

Polybrominated Biphenyls. No studies were located regarding absorption of PBBs in humans after 

inhalation exposure. However, absorption of PBBs by inhalation (and by dermal contact) in humans can 

be inferred by the relatively high levels of PBB residues detected in adipose tissue and serum of workers 

involved in PBB manufacturing (Brown et al. 1981; Landrigan et al. 1979; Stross et al. 1981). 

No studies were located regarding quantitative absorption of PBBs in animals after inhalation exposure to 

PBBs. However, increased bromine concentrations were found in the liver and adipose tissue of rats 

exposed continuously to a commercial mixture of octabromobiphenyl for 15 weeks, suggesting that 

absorption had occurred (Waritz et al. 1977). 

Polybrominated Diphenyl Ethers. No studies were located regarding absorption of PBDEs in humans 

after inhalation exposure. Evidence for the inhalation absorption of lower-brominated PBDEs in animals 

was provided by observations of systemic toxicity in rats that were intermittently exposed to a 

commercial octaBDE product (bromine content 78.7%) as dust aerosol for 13 weeks (Great Lakes 
PBBs and PBDEs 243 

5. HEALTH EFFECTS 

Chemical Corporation 2001a, 2001b). The absorption of the lower brominated BDE congeners was 

indicated by the occurrence of hepatic, thyroid, and ovarian effects in rats following exposure to 16 or 

202 mg/m3

 for 6 hours/day, 5 days/week, for 13 weeks. 

5.4.1.2 Oral Exposure 

Polybrominated Biphenyls. Quantitative oral absorption data in humans were not located, but reports of 

increased levels of PBB residues in tissues and serum of individuals accidentally exposed to contaminated 

food indicate that gastrointestinal absorption of PBBs had occurred (Eyster et al. 1983; Humphrey and 

Hayner 1975; Landrigan et al. 1979; Miceli et al. 1985; Wolff et al. 1982). 

Absorption of PBBs from the gastrointestinal tract in animals can be inferred from the numerous reports 

of adverse effects (Section 5.2.2) and increased residue levels in tissues following oral administration of 

these compounds (Section 5.4.2.2); however, few quantitative data exist. By comparing the amount of 

radioactivity in the feces of rats administered a single 1 mg/kg oral dose of 14C-2,2',4,4',5,5'-hexabromobiphenyl (BB-155) with that monitored after a single intravenous injection of the compound, it was 

estimated that ≈93% of the oral dose was absorbed over a 24-hour period (Matthews et al. 1977). Data 

obtained from similar experiments with PBB-155 later confirmed these results (Tuey and Matthews 

1980). It was also demonstrated that absorption of this hexabromobiphenyl congener was independent of 

the dose, since ≥90% was absorbed over a dose range of 1–30 mg/kg (Matthews et al. 1977). In contrast 

with the high absorption rate for the hexabromobiphenyl congener, a commercial octabromobiphenyl 

mixture (45.2% octa, 47% nona, 5.7% deca, 1.8% hepta) appeared to be less well absorbed by rats after 

administration of a single dose of 1 mg/kg (Norris et al. 1975a). Within the first 24 hours after dosing, 

61.9% of the dose was found in the feces. This indicates that at least 38.1% of the dose was absorbed, but 

absorption may have been higher, since biliary excretion may have occurred. 

Polybrominated Diphenyl Ethers. No information was located regarding absorption of PBDEs in 

humans following oral exposure. Information regarding oral absorption in animals is available from 

studies of commerical PBDE mixtures and individual PBDE congeners. As summarized below, 

absorption of decaBDE is poor, whereas the lower brominated PBDEs are readily absorbed . 

Studies with 14C-labeled decaBDE (BDE 209) in rats consistently indicate that gastrointestinal absorption 

of this congener is low at 10% or less in rats (El Dareer et al. 1987; Klasson Wehler et al. 2001; Morck 

and Klasson Wehler 2001; Morck et al. 2003; Norris et al. 1973, 1975c). Following treatment with a 
PBBs and PBDEs 244 

5. HEALTH EFFECTS 

single 1 mg/kg dose of 14C-decaBDE, administered as a former low purity commercial mixture (77.4% 

decaBDE, 21.8% nonaBDE, 0.8% octaBDE) by gavage, 90.6 and >99% of the dose was eliminated in the 

feces within 24 and 72 hours post-dosing, respectively (Norris et al. 1973, 1975b). Two feeding studies 

were conducted in which rats were exposed to a higher purity commercial mixture as unlabeled decaBDE 

(92% pure) on days 1–7 and 14C-decaBDE (98.9% pure) on day 8, followed by unlabelled decaBDE on 

days 9, 9–10 or 9–11 (El Dareer et al. 1987; NTP 1986). In the first study, dietary concentrations ranged 

from 238 to 51,100 ppm (six levels) (≈20–4,500 mg/kg/day). Recovery of radioactivity in the feces 

ranged from 91.3±4.0% to 101±4% of the amount ingested in the 72 hours following 14C-decaBDE intake 

and was not related to dose level. In the second study, rats were exposed to dietary concentrations of 277 

or 48,000 ppm (≈20 or 4,300 mg/kg/day). Recovery of radioactivity in the feces ranged from 82.5±4.7% 

to 86.4±8.5% of the dose and was not related to dose level or time of sacrifice (24, 48, or 72 hours after 

14C-decaBDE intake). For both dose levels, the percent of 14C dose remaining in the gut contents (<4%) 

and gut contents (<0.04%) decreased with time. Of the radioactivity recovered, >99% was in the feces 

and gut contents. Based on a comparison of average tissue concentrations following intravenous and oral 

administration, NTP (1986) estimated that oral absorption was 0.33±0.19% at the highest dietary level 

(50,000 ppm). 

A study of laboratory synthesized 14C-decaBDE (>98% pure) was conducted in which a single 3 μmol/kg 

(≈3 mg/kg) dose was administered to normal or bile duct-cannulated male Sprague-Dawley rats by 

gavage (Klasson Wehler et al. 2001; Morck and Klasson Wehler 2001; Morck et al. 2003). The 

compound was suspended in a novel vehicle, Lutrol F127/soya phospholipone (34:16, w/w)/water, that 

was formulated to enhance solubility and improve absorption in comparison with previous studies of 

decaBDE, including those summarized above. In the normal rats, approximately 90% (86–93%, n=8) of 

the dose was found in the feces after 3 days; cumulative recovery after 7 days was 91% (87–95%, n=4). 

For the two bile duct-cannulated rats, an average of 88 and 9.5% of the dose was recovered in the feces 

and bile, respectively, within 3 days. There was a large variation in the fecal excretion of the two 

cannulated rats, which might have been due to the fact that no bile salts were added to compensate for the 

collected bile; the investigators concluded that this likely affected absorption. Because the biliary 

excretion of decaBDE was approximately 10% of the dose, it appears that at least this much of the 

decaBDE was absorbed. The investigators could not exclude that more than 10% of the dose had been 

absorbed because 65% of the radioactivity excreted in the feces was metabolites. 

Information on oral absorption of the commercial pentaBDE mixture DE-71 and the commercial 

octaBDE mixture DE-79 is available from studies in which male Sprague-Dawley rats were fed diets 
PBBs and PBDEs 245 

5. HEALTH EFFECTS 

containing 0 or 32-33 ng/day (≈120 ng/kg/day) of either mixture in peanut oil for 21 days (Hakk et al. 

2001; Huwe et al. 2002b). The doses were designed to mimic environmental exposure levels. Liver, 

carcass, and feces were analyzed for major congeners in the penta- and octaBDE formulations 24 hours 

after the final feeding; urine was not evaluated. The study of the pentaBDE mixture assessed the 

following six congeners: 2,2’,4,4’-tetraBDE (BDE 47), 2,2’,3,4,4’-pentaBDE (BDE 85), 

2,2’,4,4’,5-pentaBDE (BDE 99), 2,2’,4,4’,6-pentaBDE (BDE 100), 2,2’,4,4’,5,5’-hexaBDE (BDE 153), 

and 2,2’,4,4’,5,6’-hexaBDE (BDE 154) (Hakk et al. 2001). Based on liver, carcass, and unrecovered 

levels of congeners, and assuming that excretion in the urine was negligible, absorption is estimated to 

have been 44.3% for penta congener BDE 85 and 84.3–92.4% for the other tetra- to hexaBDE congeners. 

The study of the octaBDE mixture assessed the following eight congeners: 2,2’,4,4’,5,5’-hexaBDE (BDE 

153), 2,2’,4,4’,5,6’-hexaBDE (BDE 154), 2,2’,3,4,4’,5’,6-heptaBDE (BDE 183), 2,3,3’,4,4’,5,6-heptaBDE (BDE 190), an unknown heptaBDE, and three unknown octaBDEs (Huwe et al. 2002b). Based on 

liver, carcass, and unrecovered levels of congeners, and assuming that excretion in the urine was 

negligible, absorption is estimated to have been 84.2–95.1% for the hexaBDEs, 68.5–79.1% for the 

heptaBDEs, and 55.7–83.3% for the octaBDEs. 

A single 14.5 mg/kg (30 μmol/kg) gavage dose of 14C-2,2',4,4'-tetraBDE (BDE 47) in corn oil was well 

absorbed by rats and mice (Örn and Klasson-Wehler 1998). Approximately 5% of the dose in rats and 

7% of the dose in mice was excreted as parent congener in the feces in 24 hours. The investigators 

concluded that these values represented the non-absorbed doses, indicating that absorption was 93–95%. 

5.4.1.3 Dermal Exposure 

Polybrominated Biphenyls. No studies were located regarding absorption of PBBs in humans or animals 

after dermal exposure to PBBs. However, absorption of PBBs through the skin in humans can be inferred 

by the relatively high levels of PBB residues detected in the adipose tissue and serum of workers involved 

in the manufacturing of these chemicals (Brown et al. 1981; Landrigan et al. 1979; Stross et al. 1981). It 

is assumed that dermal route predominates, but inhalation and/or oral exposure cannot be ruled out. 

Similarly, dermal absorption in rabbits can be inferred from reports of lethality and liver effects observed 

after application of a commercial mixture of hexabromobiphenyl to abraded and occluded dorsal skin 

(Waritz et al. 1977). 
PBBs and PBDEs 246 

5. HEALTH EFFECTS 

Polybrominated Diphenyl Ethers. No information was located regarding dermal absorption of PBDEs in 

humans. The only information regarding dermal absorption in animals is that from a study of absorption 

in an in vitro preparation (Hughes et al. 2001). In that study, 14C-decaBDE dissolved in tetrahydrofuran 

was applied to dorsal skin (three dose levels) excised from adult hairless female mice and fractions of 

receptor fluid were collected over a 24-hour period. Transfer of radioactivity to the receptor fluid was 

minimal, only 0.07 to 0.34% of the applied radioactivity. Two to 20% of the radioactivity was found in 

the skin, and the lowest dose applied had the highest percentage of the dose in the skin. Washing the skin 

with solvent 24 hours after application removed 77–92% of the applied dose. 

5.4.2 Distribution 

5.4.2.1 Inhalation Exposure 

Polybrominated Biphenyls. No studies were located regarding distribution of PBBs in humans after 

inhalation exposure. 

Limited information was located regarding distribution of PBBs in animals after inhalation exposure. 

Increased bromide concentrations were observed in the liver and adipose tissue of rats exposed 

continuously to vapors of a commercial octabromobiphenyl mixture (33% octa, 60% nona, 6% deca, 1% 

hepta) (3.5 pg octabromobiphenyl/L air at equilibrium) for 15 weeks (Waritz et al. 1977). Relative to 

controls, the concentration of bromide in liver and fat was increased by 39 and 100%, respectively; 

bromide concentration in skeletal muscle was not affected by treatment. No further details were 

provided. 

Polybrominated Diphenyl Ethers. No information was located regarding distribution of PBDEs in 

humans following inhalation exposure. 

The distribution of bromine was examined in tissues of rats after inhalation exposure octaBDE (Great 

Lakes Chemical Corporation 1978). Groups of rats were exposed to 0, 1.2, 12, 120, or 1,200 mg/m3

 of 

dusts of octaBDE 8 hours/day for 14 days. At necropsy, sections of the lungs, adipose tissue, and liver 

were collected for bromine analysis using a neutron activation technique. The results showed 

concentrations of bromine in the lungs and adipose tissue significantly higher in all groups relative to 

controls; the amounts of bromine detected were concentration-related. In the liver, the concentration of 

bromine was also elevated in all groups relative to controls except in the 1.2 mg/m3

 exposure group; the 

increases in the liver were not as marked as in the lungs or in adipose tissue. 
PBBs and PBDEs 247 

5. HEALTH EFFECTS 

5.4.2.2 Oral Exposure 

Polybrominated Biphenyls. Numerous reports have been published regarding levels of PBBs in serum, 

adipose tissue, breast milk, placenta, and cord serum of humans exposed to PBBs via the diet (Anderson 

et al. 1978d; Eyster et al. 1983; Landrigan et al. 1979; Stross et al. 1979, 1981; Wolff et al. 1979a, 1982). 

By using paired sampling, several significant correlations were determined (Eyster et al. 1983). For 

example, in parturient women from Michigan, statistically significant correlations were found between 

PBB levels in maternal serum and placenta, cord serum, breast milk, and adipose; and between PBB 

levels in adipose tissue and breast milk. In addition, there was a significant correlation between PBB 

levels in serum and feces and between serum and biliary fluid samples in farmers and chemical workers 

in Michigan. In groups of pregnant, nonpregnant, and male chemical workers the serum to adipose tissue 

PBB concentration ratios ranged from 1:140 to 1:160, but in male farmers, this ratio was 1:325–

329 (Eyster et al. 1983). The latter value is consistent with other reports regarding Michigan populations 

(Landrigan et al. 1979; Wolff et al. 1982). It is unclear why the partitioning ratios between male chemical 

workers and farmers should differ. The investigators noted that the group of farmers was much larger and 

might represent a better sample, as well as the possibilities that the farmers may have been more 

physically active or, for a variety of reasons, may have had lower total serum lipids (the amount of serum 

lipid might have affected the serum concentration of PBBs). PBB levels in body tissues and fluids are 

further discussed in Section 8.4. 

Analysis of postmortem tissue samples from 15 subjects in the Grand Rapids, Michigan area indicated 

that renal fat had the highest single PBB concentration (1.65 μg/g wet weight) and the highest mean 

concentration (0.475 μg/g) (Miceli et al. 1985). In regards to adipose, PBB concentrations in different 

tissues, could be divided into three range groups: high (ratios of 0.45–0.56, adrenal, atheromatous aorta, 

and thymus), medium (ratios of 0.1–0.28, pancreas, liver, and left ventricle), and low (ratios of 0.02–0.09, 

kidney, lung, brain, skeletal muscle, thyroid, and nonatheromatous aorta). 

As with the structurally related PCBs (Agency for Toxic Substances and Disease Registry 2000), PBBs 

are rapidly (minutes to hours) cleared from the blood and initially accumulate mainly in the liver, lungs, 

and muscle (Domino et al. 1982; Matthews et al. 1977). Due to their high affinity for lipid-rich tissues, 

PBBs are subsequently redistributed to adipose and skin for storage or metabolism in the liver, and a 

dynamic equilibrium of PBB concentrations is established among all tissues for each PBB homolog (Tuey 

and Matthews 1980). 
PBBs and PBDEs 248 

5. HEALTH EFFECTS 

In rats treated by gavage with one or four daily doses of 14C-2,2',4,4',5,5'-hexabromobiphenyl (BB 153), 

initial concentrations of radioactivity were highest in muscle, liver, and adipose tissue, but later 

redistribution to adipose tissue (4–7 days after the last dosing) resulted in lower concentrations in liver 

and muscle (Matthews et al. 1977). In rats dosed daily with 14C-BB 153 over a 30-day period, tissue 

concentrations on day 31 were (in increasing order): blood, muscle, liver, skin, and adipose and were in 

general agreement with those predicted by a physiological compartment model (Tuey and Matthews 

1980). When the model was scaled to nonlactating humans as discussed in Section 5.4.5 (Physiologically 

Based Pharmacokinetic/Pharmacodynamic Models), human intake of 9.8 g of the congener over a 

230-day period would result in peak concentrations of 720 and 2.1 ppm in fat and blood, respectively, 

5 years after the onset of exposure. The model also predicted that the body burden after 5 years would be 

5.2 g and the half-life 6.5 years. This half-life is shorter than the 12 years (median, range 4–97 years) 

calculated for hexabromobiphenyl in a Michigan cohort (Lambert et al. 1990) (see Section 5.8.1). 

In rats fed diets containing octabromobiphenyl mixture for several weeks, adipose tissue and liver 

accumulated much more bromine than did skeletal muscle (Lee et al. 1975b). For example, after 2 weeks 

of treatment, adipose of rats dosed with 50 mg/kg/day had 200 times more bromine than did adipose of 

control rats; the liver of these rats had 100 times more bromine than the livers of controls. Feeding a 

PBB-free diet for 2 weeks decreased PBB levels in liver and muscle, but not in fat. Eighteen weeks after 

exposure, the concentration of bromine in the adipose tissue of rats dosed with 50 mg/kg/day continued to 

increase to ≈840 times that of controls. Similar results were reported by Norris et al. (1975a, 1975b). 

These investigators also reported that 16 days after a single dose of octabromobiphenyl mixture in rats, 

PBB residues were present in the adrenals, adipose tissue, heart, and skin at levels ranging from 0.14 to 

0.25% of the administered dose; the liver, pancreas, and spleen contained lesser amounts. 

The distribution and elimination of PBBs from tissues were examined in rats over a period of 112 days 

after a single oral dose of FireMaster FF-1 (Domino et al. 1982). Elimination from blood was best 

described by a three-compartment model, and an elimination half-life from whole blood of 145 days was 

estimated. Relative to the three compartments (C1, C2, and C3): C1 consisted of heart, kidney, spleen, 

and whole blood; C2 included liver, lung, cerebrum, cerebellum, and testes; and C3 included 

subcutaneous fat. PBB residues in C1 rose quickly, peaked within 5 hours of dosing and then fell rapidly; 

a half-life of 3.62 hours was estimated. PBB peaked in C2 at 12 hours and then decreased; the half-life 

was 17.6 hours. In C3, levels of PBB peaked only after 1 week and remained elevated for several weeks; 

the estimated half-life was 31.1 days. Tissues with PBBs in order of increasing concentration were: 
PBBs and PBDEs 249 

5. HEALTH EFFECTS 

blood, spleen, kidney, and heart in C1, and testes, cerebrum, cerebellum, lung, and liver in C2. 

Simulations of different body fat proportions showed that reduction in body fat decreased the half-life of 

the chemical considerably. According to the investigators (Domino et al. 1982), tissues within C1 and C2 

may be at greater risk of toxicity during the subacute phase of PBB ingestion. In their view, this could 

explain the fact that blood PBB levels in Michigan families were not positively correlated with toxic 

symptoms of exposure to PBBs (see Section 5.2). 

In mink treated with FireMaster FF-1 in the diet for up to 11 months, the concentration of PBB residues 

in adipose tissue were higher than in brain and skeletal muscle at all times (Aulerich and Ringer 1979; 

Ringer et al. 1981). The source of the PBBs (FireMaster FF-1 versus food contaminated with PBBs) did 

not seem to have a significant influence on the qualitative or quantitative distribution of residues in 

tissues. Sows fed FireMaster BP-6 in the diet for 12 weeks also accumulated PBBs in adipose tissue; on a 

fat basis, the highest concentration of PBBs was found in the liver, followed by adipose, kidney, and brain 

(Werner and Sleight 1981). Distribution studies in guinea pigs after a single dose of FireMaster FF-1 

showed preferential accumulation of residues in liver, kidneys, and lungs 2 days after dosing (Ecobichon 

et al. 1983). This was followed by a slow decrease in these organs, but levels in adipose tissue reached a 

maximum between 7 and 14 days after dosing and then decreased. 

Several studies have examined the distribution of PBB residues in offspring after maternal exposure to 

PBBs during gestation and/or lactation. In 4-week-old pigs exposed in utero and via lactation to 

FireMaster BP-6, PBBs accumulated preferentially in adipose tissue and liver on a wet tissue basis. Over 

a wide range of doses, however, adipose had at least two times the PBB concentration compared to the 

liver (Werner and Sleight 1981). PBB levels in tissues of sows were comparable to those measured in 

tissues of 4-week-old pigs. On a fat basis, the liver had the highest concentration of PBBs in both sows 

and the young. In pigs exposed only in utero, PBB levels in liver and adipose were similar and 

considerably lower than in tissues of sows or 4-week-old pigs, suggesting that far more PBBs are 

transferred through lactation than through the placenta. A similar conclusion was reached in rat studies 

(McCormack and Hook 1982; Rickert et al. 1978). In contrast, PBB levels in liver and body fat of guinea 

pigs exposed briefly through lactation were considerably lower than the tissue levels acquired 

transplacentally in a 2-day period (Ecobichon et al. 1983). A biological half-life of 22 days in tissues of 

dams and pups was estimated in that study (Ecobichon et al. 1983). 

Polybrominated Diphenyl Ethers. Results of tissue distribution studies of decaBDE are consistent with 

the poor absorption and rapid fecal elimination of this congener. In male rats administered a single 
PBBs and PBDEs 250 

5. HEALTH EFFECTS 

gavage dose of 1 mg/kg of a 14C-labeled former commercial decaBDE mixture (77.4% pure containing 

21.8% nonaBDE and 0.8% octaDBE), radioactivity could be detected on day 1 in all sampled tissues 

(adipose, skin, liver, heart, adrenals, spleen, pancreas) (Norris et al. 1975b). On day 16 after dosing 

radioactivity was only detected in adrenals and spleen (0.01 and 0.06% of the administered dose per gram 

of tissue, respectively). Norris et al. (1975b) also examined the bromine content of tissues from rats 

administered 1 mg/kg/day doses of the 77.4% decaBDE mixture in the diet for 90 days, followed by an 

additional 90-day period on a control diet. During the recovery period, interim sacrifices were conducted 

on days 0, 10, 30, and 60. On recovery day 0, serum and kidneys had the same amounts of bromine as the 

controls, although adipose levels were increased approximately 4-fold and remained essentially 

unchanged over the 90-day recovery period. In the liver, the bromine content was elevated on day 0 of 

the recovery period, but decreased and stabilized near control values after recovery day 10. Tissue 

analyses in rats that were similarly exposed to 0.01-1 mg/kg/day of 77% decaBDE for 12 months showed 

adipose bromine levels that were slightly increased during the first 180 days and not significantly 

different than controls at the end of the year; no significant bromine accumulation occurred in serum, 

liver, kidneys, skeletal muscle, or testes (Norris et al. 1975b). Pregnant rats that were administered the 

77% commercial decaBDE mixture by gavage on GDs 6–15 showed no significant increase in bromine 

content in the liver of fetuses on gestation day 21 compared to controls (Norris et al. 1975b). Placental 

transfer cannot be totally ruled out since no other fetal tissues were examined. 

Two feeding studies were conducted in which rats were exposed to a higher purity commercial mixture as 

unlabeled decaBDE (92% pure) on days 1–7 and 14C-decaBDE (98.9% pure) on day 8, followed by 

unlabelled decaBDE on days 9, or 9–10, or 9–11 (El Dareer et al. 1987; NTP 1986). The compound was 

poorly absorbed in both studies. In the first study, dietary concentrations ranged from 238 to 51,100 ppm 

(≈20–4,500 mg/kg/day), and ≈91–100% of the 14C was recovered in the feces dose in 72 hours. In the 

second study, rats were exposed to dietary concentrations of 277 or 48,000 ppm (≈20 or 

4,300 mg/kg/day), and fecal recovery of 14C ranged from ≈83 to 86% of the dose. Both studies found 

only trace levels of radioactivity in any organ or tissue at any time point, and the maximum total 14C 

activity detected in the body at any time was only ≈1% of the dose. In general, the highest amounts of 14C 

were found in the liver, and there was a tendency for rats fed the lower amounts of decaBDE to retain 

more radiolabel in tissues than rats fed higher amounts. Analysis of all major organs and tissues in the 

second study found the highest levels of 14C in the gastrointestinal tract, followed by liver, kidney, lung, 

skin, and adipose. 
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5. HEALTH EFFECTS 

A study of laboratory synthesized 14C-decaBDE (>98% pure) was conducted in which a single 3 μmol/kg 

(≈3 mg/kg) dose was administered to male Sprague-Dawley rats by gavage (Klasson Wehler et al. 2001; 

Morck and Klasson Wehler 2001; Morck et al. 2003). The compound was suspended in a novel vehicle, 

Lutrol F127/soya phospholipone (34:16, w/w)/water, that was formulated to enhance solubility and 

optimize absorption. The amount of the 14C dose remaining in the body after 3 and 7 days was 

approximately 9% (calculated by subtracting total urine and feces output from 100%). The highest 

concentrations of radioactivity on a fresh weight basis were in adrenals, kidneys, heart, and liver after 

both 3 and 7 days. Based on lipid weight, plasma and liver had the highest concentrations and adipose 

tissue had the lowest concentrations at both time points. The data indicate that the highest concentrations 

were in plasma and blood-rich tissues and that decaBDE is not readily distributed to adipose. The 

investigators speculated that decaDBD does not partition into lipids and is transported through aqueous 

compartments (e.g., serum and bile) due to binding to transport proteins. As discussed below, PBDEs 

with fewer numbers of bromine preferentially accumulate in adipose. The accumulation of lower 

brominated congeners appears to be due to their partitioning and retention in lipid-rich tissues, as well as 

rates of metabolism and elimination that are likely lower than for decaBDE. 

The tissue half-life of a commercial pentaBDE product (Bromkal 70) was investigated in male and female 

Wistar rats following gavage administration of a single high dose (300 mg/kg) in corn oil (von Meyerinck 

et al. 1990). Groups of four rats/sex were sacrificed at weekly intervals until 10 weeks post-dosing for 

congeneric analysis of unspecified organs and perirenal fat. Distribution from extra-adipose tissues into 

the fat was essentially complete after 4 days. Five congeners were detected in the fat and incompletely 

characterized as a tetraBDE, two pentaBDEs, and two hexaBDEs. The half-life of the congeners 

generally increased with increasing bromination. The mean half-life of the tetraBDE congener was 

19.1 days in males and 29.9 days in females and significantly (p=0.01) different between the sexes. The 

mean half-lives of the penta- and hexaBDE congeners did not significantly differ between the sexes, 

ranging from 36.8–47.4 days for pentaBDE congener 1, 24.9–25.2 days for pentaBDE congener 2, 44.6–

55.1 days for hexaBDE congener 1, and 90.9–119.1 days for hexaBDE congener 2. Although the 

congeners were not fully identified (likely due to a lack of analytical standards), it is relevant to note that 

the predominant congeners in Bromkal 70 are 2,2’,4,4’-tetra BDE (BDE 47) and 2,2’,4,4’,5-penta BDE 

(BDE 99). 

Information on the tissue distribution of the commercial pentaBDE mixture DE-71 and the commercial 

octaBDE mixture DE-79 is available from studies in which male Sprague-Dawley rats were fed diets 

containing 0 or 32–33 ng/day (≈120 ng/kg/day) of either mixture in peanut oil for 21 days (Hakk et al. 
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5. HEALTH EFFECTS 

2001; Huwe et al. 2002b). The doses were designed to mimic environmental exposure levels. Liver, 

carcass, and feces were analyzed for major congeners in the penta- and octaBDE formulations 24 hours 

after the final feeding. The study of the pentaBDE mixture assessed the following six congeners: 

2,2’,4,4’-tetraBDE (BDE 47), 2,2’,3,4,4’-pentaBDE (BDE 85), 2,2’,4,4’,5-pentaBDE (BDE 99), 

2,2’,4,4’,6-pentaBDE (BDE 100), .2,2’,4,4’,5,5’-hexaBDE (BDE 153), and 2,2’,4,4’,5,6’-hexaBDE (BDE 

154) (Hakk et al. 2001). An average of 0.4–1.2 and 27.4–45.2% of the dosed congeners occurred in the 

liver and carcass, respectively, and 11.7–59.1% of the doses were not recovered. The congener 

distribution patterns in the liver and carcass resembled that of the commercial pentaBDE mixture, 

suggesting that there was no preferential tissue accumulation of the tested congeners. The study of the 

octaBDE mixture assessed the following eight congeners: 2,2’,4,4’,5,5’-hexaBDE (BDE 153), 

2,2’,4,4’,5,6’-hexaBDE (BDE 154), 2,2’,3,4,4’,5’,6-heptaBDE (BDE 183), 2,3,3’,4,4’,5,6-heptaBDE 

(BDE 190), an unknown heptaBDE, and three unknown octaBDEs (Huwe et al. 2002b). Four of 

congeners (BDE 90, the unknown heptaBDE, and two of the unknown octaBDEs) were not detected in 

the liver, and one congener (an unknown octaBDE) was not detected in the carcass. Of the detected 

congeners, an average of 1.1–1.7% and 16.2–62.9% of the dose occurred in the liver and carcass, 

respectively. Unlike the lower brominated congeners (tetra- to hexaBDEs) in the study of the pentaBDE 

mixture, tissue accumulation generally decreased with increasing bromination; total retentions in the liver 

and carcass were 64.4% for one of the hexaBDEs (BDE 153), 27.0–37.1% for the three heptaBDEs, and 

16.2–22.8% for two of the octaBDEs. An average of 19.6–83.3% of the congener doses were not 

recovered. The brain uptake and retention of PBDE congeners was studied in neonatal NMRI male mice 

that were given a single 1.5 mg/kg dose of 14C-2,2’,3,3’,4,4’,5,5’,6,6’-decaDBE (BDE 209), or 0.8 mg/kg 

of 14C-2,2’,4,4’,5-pentaDBE (BDE 99), by gavage in a 20% fat emulsion vehicle, on PNDs 3, 10, or 19 

and killed 24 hours or 7 days after exposure (Eriksson et al. 2002b, Viberg et al. 2001b, 2003a). In the 

study with BDE 209, after 24 hours, the mice exposed on PND 3 or 10 had 4.8 or 4.0% of the total 

administered radioactivity in the brain, respectively, whereas the brain of those exposed on PND 19 had 

only 0.6% of administered amount (Viberg et al. 2001b, 2003a). Seven days after exposure, the amount 

of radioactivity in the brain had increased approximately 2-fold in the mice exposed on PND 3 or 10 (to 

7.4 or 10.5% of the administered amount), whereas those exposed on PND 19 showed the same amount 

as at 24 hours post-exposure. These findings suggest that brain uptake of 14C-BDE 209 was more 

efficient in the younger animals and that the amount of radioactivity reaching the brain increased with 

time. The pattern of brain retention for BDE 209 was different than that for BDE 99, which showed a 

decrease over the 7-day period (Eriksson et al. 2002a). In particular, 24 hours after exposure to 

14C-BDE 209, the amount of radioactivity in the brain was between 3.7 and 5.1% of the total administered 
PBBs and PBDEs 253 

5. HEALTH EFFECTS 

amount in the three different age categories, whereas the amount declined to 1.3–2.8% of the 

administered dose after 7 days. 

Rats that were administered a single 2.2 mg (≈10 mg/kg) gavage dose of 14C-2,2',4,4'-5-pentaBDE 

(BDE 99) in corn oil and examined after 72 hours had the largest amounts of radioactivityin the carcass, 

adipose (epididymal), and gastrointestinal tract (38.8, 3.8, and 6.1% of the dose, respectively (Hakk et al. 

1999, 2002). No other tissues contained contained >1% of the radioactivity. Fractionation of the carcass 

into skin, bone, brain, eyes, and muscle showed that the majority of the 14C was in the skin. When 

deposition was expressed on a concentration basis, the highest levels of radioactivity.occurred in the most 

lipid-rich tissues, i.e., adipose, adrenals, gastrointestinal tract, and skin. 

A single 14.5 mg/kg dose of 14C-2,2',4,4'-tetraBDE (BDE 47) in corn oil was administered by gavage to 

rats and mice (Örn and Klasson-Wehler 1998). Approximately 86% of the absorbed dose in rats and 80% 

of the dose in mice remained in tissues (liver, lung, kidney, brain) at 5 days, mainly in the adipose. In 

rats, adipose tissue had an approximately 70 times higher concentration of label than the other tissues on a 

fresh weight basis and 3.5-fold higher on a lipid weight basis. The lungs had the second highest 

concentration of radiolabel with approximately twice that in the liver and kidneys. In mice, adipose 

tissues had an approximately 10-fold higher level of label than other tissues on a fresh weight basis. On a 

lipid weight basis, adipose tissue had a concentration of 14C similar to liver and twice that in lung and 

kidney. 

5.4.2.3 Dermal Exposure 

Polybrominated Biphenyls. No studies were located regarding distribution of PBBs in humans after 

dermal exposure. 

Increased liver weight and necrosis were observed in rabbits after application of an unspecified 

hexabromobiphenyl mixture to the skin, suggesting that PBBs or metabolites reached that organ (Waritz 

et al. 1977). No further information was available. 

Polybrominated Diphenyl Ethers. No studies were located regarding distribution of PBDEs in humans 

or animals after dermal exposure. 
PBBs and PBDEs 254 

5. HEALTH EFFECTS 

5.4.2.4 Other Routes of Exposure 

Polybrominated Biphenyls. In general, the distribution pattern of PBBs after parenteral administration is 

similar to that obtained after oral exposure. In rats, immediately after intravenous injection of 

14C-2,2',4,4',5,5'-hexabromobiphenyl (BB-153) adipose, skin, muscle, liver, and blood contained ≈29, 20, 

40, 10, and 1.5% of the dose, respectively (Matthews et al. 1977). Seven to 42 days postdosing, most of 

the residue in liver and muscle was redistributed to adipose tissue. The percent of the dose remaining in 

liver and muscle on day 42 was 0.8 and 3.5%, respectively. The concentration of radioactivity in skin 

remained relatively constant over a 42-day period. In a similar study in rats, the adipose/blood 

equilibrium distribution rate was found to be much higher than for any other tissue examined, and 4 days 

after dosing, adipose tissue contained ≥60% of the body burden (Tuey and Matthews 1980). 

The elimination half-times from blood and several tissues were determined in rats administered a single 

intraperitoneal dose of 10 mg/kg FireMaster BP-6 (Miceli and Marks 1981). Elimination from serum 

followed first-order kinetics, and a half-time of 23.1 weeks was calculated over a 36-week period after 

dosing. Adrenal and adipose tissue had the highest PBB concentrations at week 6, and these levels were 

maintained throughout the 36-week observation period. Concentrations of PBBs were also elevated in the 

liver, lungs, and pituitary at week 6, whereas PBB levels in brain, kidney, and spleen were several-fold 

lower. Elimination half-times from adrenal, brain, fat, liver, lung, and spleen were 43.3, 63.0, 69.3, 11.5, 

11.2, and 9.0 weeks, respectively. Elimination from heart, kidney, and pituitary did not appear to follow 

first-order kinetics; thus, elimination half-times from these tissues were not calculated. The concentration 

of PBB in adipose tissue was at least 4 times higher than in any other tissue, and unlike other tissues, 

continued to increase, reaching a maximum at week 12 postdosing. The adipose/serum ratio of PBB 

concentration increased from 222 at 6 weeks to 722 at 36 weeks, reflecting the much more rapid 

elimination of PBB from serum than from adipose tissue. The investigators estimated that, given the 

elimination half-time from fat of 69 weeks, >1 μg/g of PBB would remain in fat by the time the rats 

reached 2 years of age, the end of their lifespans. 

The distribution of PBB residues was also examined in pregnant mink and ferrets after injection of a 

mixture of 2,2',4,4',5,5'-hexabromobiphenyl (BB 153) and 2,2',3,4,4',5,5'-heptabromobiphenyl (BB 180) 

(Bleavins et al. 1980). Two hours after a single injection in the jugular vein on gestation day 37, the liver, 

kidney, and adipose tissue of ferret dams had 1.625, 0.108, and 0.124% of the dose/g tissue, respectively. 

PBB levels in fetal tissues did not exceed 0.013% of the dose (liver). In mink, PBB levels in maternal 

liver, kidney, and adipose tissue were 1.622, 0.087, and 0.031% of the dose/g tissue, respectively. Fetal 

liver had the highest amount of PBBs, 0.005% of the dose. In a different experimental series, the 
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5. HEALTH EFFECTS 

investigators (Bleavins et al. 1980) also showed that the dam's milk was the major route of offspring 

exposure since PBB levels, on a per gram or per kit basis, were significantly higher in 2-week-old kits 

than in newborn kits. The ratio of the 2-week PBB concentration to the birth concentration was 3.94/g 

and 36.66/kit. On a per kit basis, treated newborn kits accumulated 0.80% of the maternal dose through 

in utero exposure. 

Polybrominated Diphenyl Ethers. No relevant information was located regarding distribution of PBDEs 

following exposure by non-natural routes of administration. 

5.4.3 Metabolism 

Polybrominated Biphenyls. Information regarding the metabolism of PBBs in humans is limited. 

Chromatographic analysis of serum samples from Michigan dairy farmers and from Michigan Chemical 

Corporation employees revealed some differences in peak profile between these two groups and between 

these two groups and the peak profile of FireMaster BP-6 (Wolff and Aubrey 1978; Wolff et al. 1979a). 

The concentration of two pentabromobiphenyls was lower in the farmers than in the chemical workers. 

Both farmers and workers had a significantly lower amount of 2,2',3,4,4',5,5'-heptabromobiphenyl than 

FireMaster BP-6. Other minor differences between the groups were also apparent. The differences in 

peak profiles between farmers and chemical workers were attributed to different routes of exposure. 

Farmers had predominantly dietary exposure to PBBs which, according to the authors (Wolff and Aubrey 

1978), could have undergone partial metabolism in the animal food source (see below). It should be 

noted that chemical transformation of the PBBs due to cooking of meat or pasteurization of milk whould 

not be expected since the temperatures reached during these processes is probably not high enough. As 

discussed in Chapter 6, temperatures must exceed ≈500 °C for structural alterations of PBBs to occur. 

Nevertheless, a significant reduction (36–52%) in the concentration of PBBs in pressure-cooked meat 

relative to raw meat (due to loss of fat) has been reported (Zabik et al. 1978). The decreased heptabromobiphenyl peak in farmers and workers relative to FireMaster BP-6 may reflect poor absorption of this 

congener since it is not expected to be metabolized readily (Wolff and Aubrey 1978; Wolff et al. 1979a). 

Human exposure to PBBs in the Michigan contamination episode occurred primarily through 

consumption of contaminated meat and dairy products. The limited information available regarding the 

metabolism of PBBs in dairy cattle is insufficient to ascertain whether humans ingested PBBs or 

metabolic products of PBBs. In a controlled study, cows fed single or repeated doses of FireMaster BP-6 

excreted 50% of the dose in the feces as parent compound (Willet and Durst 1978). Tissues, feces, or 
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5. HEALTH EFFECTS 

urine were not analyzed for metabolites. Results of studies in rats, and also in vitro studies (see below), 

have shown that highly brominated PBB congeners, such as the major components of FireMaster BP-6, 

undergo little or no metabolic transformation. Based on the existing information, it seems reasonable to 

assume that in the Michigan contamination episode, humans consumed mainly unchanged penta-, hexa-, 

and heptabromobiphenyls. 

The in vivo metabolism of some PBB congeners and of commercial PBB mixtures has been investigated 

in a limited number of animal studies. For example, in pigs, intraperitoneal injection of 4-bromobiphenyl 

yielded three urinary metabolites: 4'-bromo-4-biphenylol (3% of the dose), bromobiphenylol (traces), and 

4'-bromobiphenylol (0.5% of the dose) (Kohli and Safe 1976). 4,4'-Dibromobiphenyl (BB 15) yielded 

four urinary metabolites: 4,4'-dibromo-3-biphenylol (5% of the dose), 3,4'-dibromo-4-biphenylol (1% of 

the dose), 4'-bromo-3-methoxy-4-biphenylol (1% of the dose), and traces of a dibromomethoxybiphenyl. 

The authors suggested these results indicate that metabolism of BB 15 occurs through the formation of an 

arene oxide. The major urinary metabolite of FireMaster BP-6 was a pentabromobiphenylol (1% of the 

dose), which could have resulted from direct hydroxylation of the minor pentabromobiphenylol isomers 

in FireMaster BP-6 or by debromination/hydroxylation of the major congener, 2,2',4,4',5,5'-hexabromobiphenyl (BB 153). 

In rabbits, metabolism of 2-bromobiphenyl yielded two polar metabolites, one metabolite was identified 

as 2'-bromo-4-biphenylol (1% of the dose), and the other metabolite (traces) was also a monohydroxylated derivative, but the position of the hydroxyl group was not determined (Kohli et al. 1978). 

3-Bromobiphenyl produced a major metabolite (4% of the dose) identified as either 3-bromo-4-biphenylol 

or 5-bromo-2-biphenylol; a minor dihydroxylated metabolite was also detected. 4-Bromobiphenyl 

yielded two metabolites: 4'-bromo-4-biphenylol (2% of the dose) and 4'-bromo-3,4-biphenyldiol (1.5% 

of the dose). Experiments with tritiated 4-bromobiphenyl suggest that the metabolism of this congener 

involves the formation of an arene oxide. 

Similar results have been reported in rats (Sparling et al. 1980). 4'-Bromo-4-biphenylol was the major 

metabolite of 4-bromobiphenyl (BB 3). 2-Bromobiphenyl (BB 1) was metabolized to 2-bromo-4,4'-biphenyldiol and 2-bromo-4',5-biphenyldiol; 2-bromo-5-biphenylol was a minor metabolite. 

3-Bromobiphenyl (BB 2) also yielded diols as major metabolites: 3-bromo-4,4'-biphenyldiol and an 

unknown diol. The main conclusions of this experiment were: the major site of hydroxylation is at the 

para position of the unsubstituted phenyl ring, and also at the para position of the ring for BB 1 and 

BB 2; substitutions in positions 2 and 3 tend to direct hydroxylation to position para and ortho (minor) to 
PBBs and PBDEs 257 

5. HEALTH EFFECTS 

the substituents; and the 2- and 3-hydroxylated products are subsequently dehydroxylated, whereas the 

4'-hydroxy congener is not. 

In contrast to the lower brominated congeners, no major metabolites were identified in the urine or feces 

of rats treated with a single intraperitoneal dose of 2,2'4,4'5,5'-hexabromobiphenyl, suggesting that this 

congener is stable and persistent (Safe et al. 1978). Analyses of the feces of dogs administered 

FireMaster BP-6 orally revealed the presence of a metabolite identified as 6-hydroxy-2,2',4,4',5,5'-hexabromobiphenyl (Gardner et al. 1979). This metabolite was not found in the liver, but the parent 

compound was identified. Since hydroxylation in position 6 of highly substituted congeners is unlikely, it 

was postulated that the metabolite found in the feces was formed by microbial metabolism of the PBB in 

the intestinal tract. The in vitro metabolism of 2,2',4,4',5,5'-hexabromobiphenyl (BB 153) with liver 

microsomes of rats induced with either BB 153 or FireMaster BP-6 produced three major metabolic 

fractions: lipophilic ether soluble polar metabolites, trichloroacetic acid (TCA) soluble conjugates, and 

macromolecular adducts (Purdy and Safe 1980). 

The NADPH-dependent metabolism of a PBB mixture was studied in vitro with liver microsomes of rats 

induced with PB, PBB, or 3-MC (Dannan et al. 1978a). Of the 12 major components of the mixture, only 

2,2',4,5,5'-pentabromobiphenyl (BB 101) and a hexabromobiphenyl were metabolized by microsomes 

from PB- or PBB-treated rats. Of seven structurally identified PBB components, only BB 101 had a 

bromine-free para position. Although BB 101, 2,3',4,4',5-pentabromobiphenyl (BB 118), and 

2,2',3,4,4',5'-hexabromobiphenyl (BB 138) have two adjacent unsubstituted carbons, only BB 101 was 

metabolized. No significant metabolism occurred when the PBB mixture was incubated with microsomes 

of control rats or MC-induced rats. When 2,2'-dibromobiphenyl (BB 4) and 4,4'-dibromobiphenyl 

(BB 15) were incubated with liver microsomes of PB-treated rats, only BB 4 was metabolized. These 

results suggest that the presence of a free para position is required for the metabolism of brominated 

biphenyls and that the bromine content of the molecule is less important in determining metabolism than 

the position of bromines on the biphenyl nucleus. 

A more recent study with hepatic microsomes of induced rats showed that MC pretreatment increased the 

NADPH-dependent metabolism of PBB congeners (di-, tri-, and tetrabrominated), which had adjacent 

unsubstituted ortho and meta positions on at least one ring (Mills et al. 1985). Some penta- and 

hexabromobiphenyls that have adjacent unsubstituted ortho and meta positions were not metabolized, 

suggesting that further bromination prevents metabolism. Pretreatment with PB increased the microsomal 

metabolism of congeners that have adjacent unsubstituted meta and para positions on at least one ring. It 
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5. HEALTH EFFECTS 

was concluded that the rates of metabolism of PBB congeners depends on the position of the bromines 

and the form of the cytochrome P-450 induced. The ability to metabolize PBBs also depends on the 

species. For example, hepatic microsomes isolated from rats have a greater potential to metabolize PBBs 

than hepatic microsomes isolated from pigeons (Borlakoglu and Wilkins 1993). 

Polybrominated Diphenyl Ethers. DecaBDE (BDE 209) is predominantly excreted in the feces as 

unabsorbed parent molecule. Data are available indicating that the small amount of decaBDE that is 

absorbed can be metabolized (El Dareer et al. 1987; Morck et al. 2003; NTP 1986). A feeding study was 

conducted in which rats were exposed to a high purity mixture as unlabeled decaBDE (92% pure) on days 

1–7 and 14C-decaBDE (98.9% pure) on day 8, followed by unlabelled decaBDE on days 9, or 9–10, or 9–

11, at dietary concentrations of 277 or 48,000 ppm (El Dareer et al. 1987; NTP 1986). Recovery of 

radioactivity in the feces ranged from 82.5±4.7 to 86.4±8.5% of the dose, and was not related to decaBDE 

dose level or time of sacrifice (24, 48, or 72 hours after consumption of 14C-decaBDE). For both dose 

levels, the percent of 14C dose remaining in the gut contents (<4%) and gut contents (<0.04%) decreased 

with time. Of the radioactivity recovered, >99% was in the feces and gut contents. HPLC and UV 

spectral analyses of extracts of feces collected on days 9–11 indicated the presence of decaBDE (81% of 

recovered radioactivity) and three main unidentified metabolites. The total radioactivity present as the 

metabolites tended to increase with increasing dietary level of decaBDE (1.5 and 27.9% of total recovery 

at the low and high dose, respectively). Since absorption of decaBDE was minimal, El Dareer et al. 

(1987) speculated that the greater extent of metabolism in the high dose rats may be due to induction of 

hepatic metabolizing enzymes, or gastrointestinal tract bacteria, during the period of feeding unlabeled 

decaBDE. Analysis of feces from rats following a single 1.07 mg/kg intravenous dose of 14C-decaBDE 

similarly indicated that the excreted material was predominantly unchanged compound and three main 

unidentified metabolites (El Dareer et al. 1987; NTP 1986). Unchanged decaBDE comprised 36.5 and 

40.4% of the totals for the 0–48-hour and 48–72-hour collection periods, respectively. The HPLC 

retention times of the metabolites were similar in the oral and intravenous studies. A single metabolite 

was detected in the bile of rats following a single 0.9 mg/kg intravenous dose of 14C-decaBDE (El Dareer 

et al. 1987). Approximately 7% of the administered dose appeared in the bile in 4 hours, and <1% of this 

amount was unchanged decaBDE. 

A study of laboratory synthesized 14C-decaBDE (>98% pure) was conducted in which a single 3 μmol/kg 

(≈3 mg/kg) dose was administered to normal or bile duct-cannulated male Sprague-Dawley rats by 

gavage (Morck and Klasson Wehler 2001; Morck et al. 2003). The compound was emulsified in a novel 

vehicle, Lutrol F127/soya phospholipone (34:16, w/w) in water, that was formulated to enhance solubility 
PBBs and PBDEs 259 

5. HEALTH EFFECTS 

and optimize absorption. Feces, bile, and several tissues (liver, kidney, lung, small intestine wall, and 

adipose) were qualitatively analyzed for metabolites using gel permeation chromatography (GPC) and gas 

chromatography-mass spectrometry (GC-MS). In the normal rats, approximately 90% of the dose was 

excreted in the feces in 3 days, mainly as metabolites (65% of dose). In the the two bile duct-cannulated 

rats, an average of 9.5% of the dose was excreted via the bile in 3 days, almost all of which represented 

metabolites. Metabolites were characterized as nonextractable, water-soluble, lipid-bound, phenolic 

metabolites, and parent compound/neutral metabolites. Feces from normal rats showed formation of 

phenolic and neutral metabolites. Metabolites in the nonconjugated phenolic fraction included six pentato heptaBDEs containing a guaiacol structure (a hydroxy and a methoxy group) on one of the rings 

(proposed to be on vicinal carbons), as well as traces of monohydroxylated diphenyl ethers with at least 

six bromine atoms and uncharacterized metabolites having eight bromine atoms. DecaBDE was the main 

component in the neutral fraction of the feces, but trace amounts (<0.5% of the decaBDE) of three 

nonaBDEs also formed. The mechanism by which decaBDE was debrominated to the methoxyhydroxylated penta- to hepatBDEs was not clear, but was inferred to include a catechol (dihydroxy 

compound) and/or other reactive intermediates. Analysis of bile showed metabolites that were mainly 

lipid-bound on day 1 and subsequently water soluble. Neutral compounds in the bile consisted mainly of 

parent decaBDE and traces of the three nonaBDEs that were found in the feces. Phenolic metabolites in 

the bile included eight compounds that were the same as those found in the feces. The tissue evaluations 

showed a high concentration of radioactivity in the liver and small intestine wall (27 and 61%, 

respectively) that was largely nonextractable, and therefore was assumed to be bound covalently to 

macromolecules and indicative of adducts formed by metabolism via reactive metabolites. The 

metabolites identified in the feces of the rats were probably due to debromination of decaBDE, followed 

by metabolic substitution of vicinal carbons by methoxy or hydroxy groups. The authors did not rule out 

that the intestinal endoderm or gut microflora could play a role in some of this metabolism. PBDEs with 

fewer bromines probably have lower rates of metabolism (and elimination) than decaBDE because lower 

brominated PBDEs partition into, and are retained in, lipid-rich tissues, whereas decaBDE does not 

partition into lipids. 

Information on the metabolism of the commercial pentaBDE mixture DE-71 and the commercial 

octaBDE mixture DE-79 is available from studies in which male Sprague-Dawley rats were fed diets 

containing 0 or 32–33 ng/day (≈120 ng/kg/day) of either mixture in peanut oil for 21 days (Hakk et al. 

2001; Huwe et al. 2002b). The doses were designed to mimic environmental exposure levels. The 

predominant congeners in the pentaBDE formulation were 2,2’,4,4’-tetraBDE (BDE 47) and 

2,2’,4,4’,5-pentaBDE (BDE 99). The octaBDE formulation was mainly comprised of hexaBDE through 
PBBs and PBDEs 260 

5. HEALTH EFFECTS 

nonaBDE congeners and only contained tetra- and pentaBDEs as minor components. Liver, carcass, and 

feces were analyzed for major congeners in the penta- and octaBDE formulations 24 hours after the final 

feeding. Urine was not analyzed because the lipophilic nature of the compounds and previous studies 

indicated that PBDEs are not excreted in the urine. 

The study of the pentaBDE mixture assessed the following six congeners: 2,2’,4,4’-tetraBDE (BDE 47), 

2,2’,3,4,4’-pentaBDE (BDE 85), 2,2’,4,4’,5-pentaBDE (BDE 99), 2,2’,4,4’,6-pentaBDE (BDE 100), 

2,2’,4,4’,5,5’-hexaBDE (BDE 153), and 2,2’,4,4’,5,6’-hexaBDE (BDE 154) (Hakk et al. 2001). An 

average of 0.4–1.2 and 27.4–45.2% of the dosed congeners occurred in the liver and carcass, respectively. 

The congener distribution patterns in the liver and carcass resembled that of the commercial pentaBDE 

mixture, suggesting that there was no preferential tissue accumulation of the tested congeners. An 

average of 11.7–59.1% of the congener doses was not recovered, suggesting that significant metabolism 

of some of the congeners occurred. The study of the octaBDE mixture assessed the following eight 

congeners: 2,2’,4,4’,5,5’-hexaBDE (BDE 153), 2,2’,4,4’,5,6’-hexaBDE (BDE 154), 2,2’,3,4,4’,5’,6-

heptaBDE (BDE 183), 2,3,3’,4,4’,5,6-heptaBDE (BDE 190), an unknown heptaBDE, and three unknown 

octaBDEs (Huwe et al. 2002b). Four of congeners (BDE 90, the unknown heptaBDE, and two of the 

unknown octaBDEs) were not detected in the liver, and one congener (an unknown octaBDE) was not 

detected in the carcass. Of the detected congeners, an average of 1.1–1.7% and 16.2–62.9% of the dose 

occurred in the liver and carcass, respectively. Unlike the lower brominated congeners (tetra- to 

hexaBDEs) in the study of the pentaBDE mixture, tissue accumulation generally decreased with 

increasing bromination; total retentions in the liver and carcass were 64.4% for one of the hexaBDEs 

(BDE 153), 27.0–37.1% for the three heptaBDEs, and 16.2–22.8% for two of the octaBDEs. An average 

of 19.6–83.3% of the congener doses were not recovered, indicating that most of the congeners were 

metabolized to some extent 

The metabolism of 14C-2,2',4,4'-5-pentaBDE (BDE 99) was investigated in normal and bile ductcannulated rats that were administered a single 2.2 mg (≈10 mg/kg) dose in corn oil by gavage (Hakk et 

al. 1999, 2002). The feces was the major route of elimination as shown by fecal recovery of 43 and 86% 

of the administered 14C in 72 hours in the normal and bile-duct cannulated rats, respectively. The feces 

from the normal rats contained predominantly parent compound and minor amounts of metabolites 

(>90 and <10% of the extracted radioactivity, respectively). The fecal metabolites were incompletely 

identified as two mono-OH-pentaBDEs and two mono-OH-tetraBDEs, indicating that some 

debromination occurred. Metabolites found in the bile included two mono-OH-pentaBDEs, three di-OHpentaBDEs, and two possible thiol-substituted pentaBDEs. The possible thiol-substituted pentaBDEs 
PBBs and PBDEs 261 

5. HEALTH EFFECTS 

could have been formed through the mercapturic acid pathway, but no glucuronide or sulfate conjugates 

were found in the bile. Evidence for reactive intermediates in the feces of normal rats was indicated by 

high nonextractable fractions ranging from 18–52%. Cumulative excretion of radiolabel in the urine at 

72 hours, consisting exclusively of uncharacterized metabolites, was 1 and 0.3% of the dose in the normal 

and bile duct-cannulated rats, respectively. Absorbed BDE 99 was preferentially distributed to lipophilic 

tissues, in which it accumulated as parent compound; the only tissue in which metabolites were detected 

was the liver, which contained only trace amounts of OH-metabolites. 

Rats given a single gavage dose of 14.5 mg/kg dose of 14C-2,2',4,4'-tetraBDE (BDE 47) in corn oil 

excreted 14% of the dose in the feces and <0.5% in the urine over a 5-day period (Örn and KlassonWehler 1998). The majority of the fecal 14C (≈85% of the parent compound/metabolite fraction) was 

parent compound. Six metabolites were detected in the feces, but were not precisely identified; the 

metabolites were tentatively characterized as hydroxlated derivatives (two ortho-, one meta-, and two 

para-OH-tetraBDEs) and a trace amount of a thiol-tetraBDE. Mice treated in the same manner excreted 

20% of the dose in the feces and 33% in the urine over 5 days, indicating that the mouse is more capable 

of metabolizing BDE 47 than rats. The majority of the fecal 14C in mice (≈70% of the parent 

compound/metabolite fraction) was the metabolite fraction, which contained the same six metabolites 

characterized in the rat feces. Approximately 20% of the mouse urinary 14C was characterized as parent 

compound; the remainder could not be identified, but was speculated to arise from decomposition of a 

labile metabolite(s). Unchanged tetraBDE was the major compound in all tissues analyzed in both 

species; minor amounts of hydroxylated tetraBDE metabolites were also detected. A water-soluble 

metabolite was detected in urine from mice, but could not be isolated. 

Following oral exposure of male Sprague-Dawley rats to decaBDE (BDE-209), 13 phenolic metabolites 

were determined in the plasma (Sandholm et al. 2003). The major metabolites were characterized as a 

hydroxyl-octaBDE, a hydroxyl-nonaBDE, and a hydroxyl-methoxy-hexaBDE. In addition to the 

debromination reactions, the presence of a methoxy group is suggestive of methylation, and possibly 

other metabolic processes, by bacteria of the gut. 

5.4.4 Elimination and Excretion 

5.4.4.1 Inhalation Exposure 

No studies were located regarding excretion of PBBs or PBDEs in humans or animals after inhalation 

exposure. 
PBBs and PBDEs 262 

5. HEALTH EFFECTS 

5.4.4.2 Oral Exposure 

Polybrominated Biphenyls. No studies were located that provide information on percentage of ingested 

PBBs excreted by humans. However, PBBs in biliary fluid of a group of farmers and chemical workers 

ranged from undetected to 70 μg/L, and the correlation between serum PBB levels and levels in bile was 

statistically significant (Eyster et al. 1983). Similarly, PBB levels in feces ranged from undetected to 

862 μg/kg, and the correlation between serum PBB levels and fecal levels was also statistically significant 

(Eyster et al. 1983). 

Serum half-life values have been estimated using human data from the Michigan PBB cohort (Blanck et 

al. 2000b; Lambert et al. 1990; Rosen et al. 1995). A median half-life of 12.0 years (95% CI 4–97 years) 

was estimated based on two serum measurements from 15 women (≥20 years of age) with an initial serum 

PBB level of ≥5 ppb (Lambert et al. 1990). An analysis of 51 women (≥18.8 years of age) and 112 men 

(≥18.1 years of age) with at least two measurements 1 year apart and an initial PBB level of ≥20 ppb 

found a median half-life of 13.0 years (95% CI 6.3–infinite years) and 10.0 years (95% CI 6.7–

20.0 years), respectively (Rosen et al. 1995). Based on a median half-life of 10.8 years (95% CI 9.2–

14.7 years) for the entire group (163 persons, median PBB level 45.5 ppb), it was estimated that it will 

take more than 60 years for their PBB levels to fall below a detection limit of 1 ppb. 

Determinants of PBB serum decay were investigated in 380 Michigan women (≥16 years of age) who had 

an initial PBB level of at least 2 ppb and at least two measurements taken when they were not pregnant 

(Blanck et al. 2000b). The mean initial PBB level was 20.9 ppb (standard deviation 78.7), and the mean 

time between the first and last measurement was 4.2 years (range 16.0–75.2 years). A total of 109 women 

(29%) did not have a reduction in serum PBBs over time. Assuming that PBBs reached equilibrium in 

the body before substantial amounts were eliminated and before the first serum measurements were taken, 

the entire body was modeled as a single compartment for PBBs with exponential decay. The median 

PBB half-life in the entire group was 13.5 years (95% CI 10.5–23.2 years). Subject-specific decay rate 

estimates were regressed on predictor variables including initial age, body mass index (BMI), smoking 

history, breast-feeding duration, and parity. The serum PBB decay rate was slower, resulting in a longer 

half-life, with higher initial PBB levels; women with initial PBB levels of <10 and >10 ppb had median 

half-lives of 12.9 and 28.7 years, respectively. The PBB decay rate was also slower (p=0.03) in women 

with an initial BMI above the median (BMI≥23). Increasing number of pregnancies between the first and 

last measurement was also associated with slower decay, but the effect was of borderline statistical 
PBBs and PBDEs 263 

5. HEALTH EFFECTS 

significance (p=0.06). Breast feeding as either a continuous variable or as categorized by tertiles of 

duration (<3, 3–9, or >9 months), age, and smoking were not associated with serum PBB decay. 

Lactation constitutes the most important route of excretion of PBB in lactating women. Numerous 

studies reported PBB levels in breast milk from Michigan women (Brilliant et al. 1978; Eyster et al. 1983; 

Humphrey and Hayner 1975; Jacobson et al. 1984; Landrigan et al. 1979). PBB levels in breast milk on a 

lipid basis ranged from undetected to 92,667 μg/kg, with a median of 250 μg/kg, in a group of parturient 

women from Michigan (Eyster et al. 1983). Regression analysis of the data revealed that on a lipid basis, 

PBBs are 107–119 times more concentrated in milk than in serum. Also, adipose PBB levels are 1.1–

1.5 times higher than the breast milk levels when milk levels were ≥100 μg/kg. 

The importance of PBB transfer through lactation in experimental animals is discussed in Section 5.4.2.2. 

There is limited information regarding excretion of PBBs in experimental animals. Rats dosed once with 

14C-2,2',4,4',5,5'-hexabromobiphenyl (BB 153) excreted 7.9% of the dose in the feces within the first 

24 hours; urinary excretion data were not provided (Matthews et al. 1977). It was estimated that <10% of 

the administered dose would ever be excreted. These results are consistent with those of other 

investigators who report that this congener is stable and persistent (Safe et al. 1978). In rats gavaged with 

14C-2,2',4,4',5,5'-hexabromobiphenyl for 22 days, between 10 and 20% of the daily dose was excreted 

daily in the feces; this value was predominantly due to elimination of unabsorbed PBB (Tuey and 

Matthews 1980). In monkeys, the main route of excretion of hexabromobiphenyl residues was also in the 

feces (Rozman et al. 1982). Between 60 and 70% of the administered dose was excreted in the feces in 

the first 11 days after dosing; urinary excretion was minimal. The difference between the absorption rate 

reported by Matthews et al. (1977) and that reported by Rozman et al. (1982) can probably be accounted 

for by differences in the experimental designs. 

Rats treated with a single gavage dose of 14C-octabromobiphenyl excreted <1% of the administered dose 

in urine and expired air over a 16-day period (Norris et al. 1975a). Within the first 24 hours after dosing, 

61.9% of the dose was present in the feces. The proportion that represents unabsorbed PBB is not known. 

By day 16, 74% of the administered dose had been recovered in the feces. 

Polybrominated Diphenyl Ethers. Oral studies with 14C-labeled decaBDE consistently show that this 

PBDE is rapidly and predominantly eliminated in the feces (El Dareer et al. 1987; Klasson Wehler et al. 

2001, Morck and Klasson Wehler 2001; Morck et al. 2003; Norris et al. 1973, 1975c). In rats treated by 
PBBs and PBDEs 264 

5. HEALTH EFFECTS 

gavage with a single 1 mg/kg dose of a low purity former commercial mixture (77.4% 14C-decaBDE, 

21.8% nonaBDE, 0.8% octaBDE) in corn oil, most of the excreted radioactivity measured over a 16-day 

period was in the feces and <1% of the label was detected in the urine and expired air (Norris et al. 1973, 

1975c). No significant differences in excretion patterns were observed between males and females. 

Approximately 91% of the decaBDE-derived radioactivity was found in the feces in the first 24 hours and 

almost all radioactivity (>99%) was accounted for by day 2 after dosing. Two feeding studies were 

conducted in which rats were exposed to a higher purity commercial mixture as unlabeled decaBDE (92% 

pure) on days 1–7 and 14C-decaBDE (98.9% pure) on day 8, followed by unlabelled decaBDE on days 9, 

or 9–10, or 9–11 (El Dareer et al. 1987; NTP 1986). In the first study, dietary concentrations ranged from 

238–51,100 ppm (six levels) (≈20–4600 mg/kg/day). Recovery of radioactivity in the feces ranged from 

91.3±4.0 to 101±4% of the amount ingested in the 72 hours following 14C-decaBDE intake and was not 

related to dose level. In the second study, rats were exposed to dietary concentrations of 277 or 

48,000 ppm (≈25 or 4,300 mg/kg/day). Recovery of radioactivity in the feces ranged from 82.5±4.7 to 

86.4±8.5% of the dose and was not related to dose level or time of sacrifice (24, 48, or 72 hours after 

14C-decaBDE intake). For both dose levels, the percent of 14C dose remaining in the gut contents (<4%) 

and gut contents (<0.04%) decreased with time. Of the radioactivity recovered, >99% was in the feces 

and gut contents. Urinary excretion of 14C accounted for ≤0.01% of the dose. 

In a study of laboratory synthesized 14C-decaBDE (>98% pure), a single 3 μmol/kg (≈3 mg/kg) dose was 

administered to normal or bile duct-cannulated male Sprague-Dawley rats by gavage (Klasson Wehler et 

al. 2001; Morck and Klasson Wehler 2001; Morck et al. 2003). The compound was suspended in a novel 

vehicle, Lutrol F127/soya phospholipone (34:16, w/w)/water, that was formulated to enhance solubility 

and optimize absorption. In the normal rats, approximately 90% (86–93%, n=8) of the dose was found in 

the feces after 3 days; cumulative recovery after 7 days was 91% (87-95%, n=4). For the two bile ductcannulated rats, an average of 88 and 9.5% of the dose was recovered in the feces and bile, respectively, 

within 3 days. There was a large variation in the fecal excretion of the two cannulated rats that might 

have been due to the fact that no bile salts were added to compensate for the collected bile. Urinary 

excretion of 14C was negligible for all groups at <0.1% of the dose. DecaBDE was speculated to have 

higher rates of metabolism and elimination than PBDEs with fewer numbers of bromines because it is 

unlikely to partition into lipids and is transported through aqueous compartments (e.g., serum and bile), 

whereas lower brominated congeners preferentially accumulate in the body due to their partitioning and 

retention in lipid-rich tissues. 
PBBs and PBDEs 265 

5. HEALTH EFFECTS 

Information on oral absorption of the commercial pentaBDE mixture DE-71 and the commercial 

octaBDE mixture DE-79 is available from studies in which male Sprague-Dawley rats were fed diets 

containing 0 or 32–33 ng/day (≈120 ng/kg/day) of either mixture in peanut oil for 21 days (Hakk et al. 

2001; Huwe et al. 2002b). The doses were designed to mimic environmental exposure levels. Liver, 

carcass, and feces were analyzed for major congeners in the penta- and octaBDE formulations 24 hours 

after the final feeding; urine was not evaluated. The study of the pentaBDE mixture assessed the 

following six congeners: 2,2’,4,4’-tetraBDE (BDE 47), 2,2’,3,4,4’-pentaBDE (BDE 85), 

2,2’,4,4’,5-pentaBDE (BDE 99), 2,2’,4,4’,6-pentaBDE (BDE 100), 2,2’,4,4’,5,5’-hexaBDE (BDE 153), 

and 2,2’,4,4’,5,6’-hexaBDE (BDE 154) (Hakk et al. 2001). One of the penta congeners, BDE 85, 

occurred at a relatively high level in the feces (55.8% of the adminstered amount). Fecal excretion of the 

five other tetra- to hexaBDE congeners ranged from 7.6 to 15.8% of the dose. The study of the octaBDE 

mixture assessed the following eight congeners: 2,2’,4,4’,5,5’-hexaBDE (BDE 153), 

2,2’,4,4’,5,6’-hexaBDE (BDE 154), 2,2’,3,4,4’,5’,6-heptaBDE (BDE 183), 2,3,3’,4,4’,5,6-heptaBDE 

(BDE 190), an unknown heptaBDE, and three unknown octaBDEs (Huwe et al. 2002b). Fecal excretion 

ranged from 4.9 to 15.9% of the dose for the hexaBDE congeners, 20.9–31.5% of the dose for the 

heptaBDE congeners, and 16.7–44.3% of the dose for the octaBDE congeners. 

The elimination of 14C-2,2',4,4'-5-pentaBDE (BDE 99) was investigated in normal and bile ductcannulated rats that were administered a single 2.2 mg (≈10 mg/kg) dose in corn oil by gavage (Hakk et 

al. 1999, 2002). The main route of elimination was via the feces as shown by cumulative fecal recovery 

of 43 and 86% of the administered radiolabel in 72 hours in the normal and bile-duct cannulated rats, 

respectively. The difference in the excreted amounts of radioactivity suggests that bile salts are needed 

for the intestinal uptake of BDE 99. Cumulative urinary excretion of radiolabel over 72 hours was ≈1% 

and ≈0.3% of the dose in the normal and bile duct-cannulated rats, respectively. The cumulative biliary 

excretion of radiolabel in cannulated rats was ≈4%. 

Rats given a single gavage dose of 14.5 mg/kg dose of 14C-2,2',4,4'-tetraBDE (BDE 47) in corn oil 

excreted 14% of the dose in the feces and <0.5% in the urine over a five day period (Örn and KlassonWehler 1998). Mice treated in the same manner excreted 20% of the dose in the feces and 33% in the 

urine over the same time period. The amount of non-absorbed material in the feces on day 1 was 

approximately 5 and 7% of the dose in rats and mice, respectively, suggesting efficient absorption in both 

species (Örn and Klasson-Wehler 1998). 
PBBs and PBDEs 266 

5. HEALTH EFFECTS 

5.4.4.3 Dermal Exposure 

No studies were located regarding excretion of PBBs or PBDEs in humans or animals after dermal 

exposure. 

5.4.4.4 Other Routes of Exposure 

Polybrominated Biphenyls. Rats given a single intravenous dose of 14C-2,2',4,4',5,5'-hexabromobiphenyl 

(BDE 153) excreted a cumulative 0.96, 3.3, and 6.6% of the dose in the feces 1, 7, and 42 days after 

dosing, respectively (Matthews et al. 1977). Only traces (0.1% of the dose) were excreted in the urine. 

Two decay components were calculated from excretion data; an initial decay rate of 1.05% of the 

dose/day and a later rate of 0.15% of the dose/day. Biliary excretion accounted for 0.68% of the dose 

between 0 and 4 hours after dosing. Analysis of bile and feces showed that at least 95% of the 

radioactivity corresponded to the parent compound. Moreover, in rats, ≈35% of the radioactivity excreted 

in the bile during the first week after a single dosing was reabsorbed (Tuey and Matthews 1980). 

Parenteral administration of mono- and dibromobiphenyls to rats, rabbits, and pigs suggests that the urine 

is an important route of excretion for polar metabolites (Kohli and Safe 1976; Kohli et al. 1978; Sparling 

et al. 1980). However, cumulative urinary excretion did not account for more than 5% of the 

administered doses. Data regarding fecal excretion were not provided. 

Polybrominated Diphenyl Ethers. No relevant information was located regarding elimination of PBDEs 

following exposure by non-natural routes of administration. 

5.4.5 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) Models 

Physiologically based pharmacokinetic (PBPK) models use mathematical descriptions of the uptake and 

disposition of chemical substances to quantitatively describe the relationships among critical biological 

processes (Krishnan et al. 1994). PBPK models are also called biologically based tissue dosimetry 

models. PBPK models are increasingly used in risk assessments, primarily to predict the concentration of 

potentially toxic moieties of a chemical that will be delivered to any given target tissue following various 

combinations of route, dose level, and test species (Clewell and Andersen 1985). Physiologically based 

pharmacodynamic (PBPD) models use mathematical descriptions of the dose-response function to 

quantitatively describe the relationship between target tissue dose and toxic end points. 
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5. HEALTH EFFECTS 

PBPK/PD models refine our understanding of complex quantitative dose behaviors by helping to 

delineate and characterize the relationships between: (1) the external/exposure concentration and target 

tissue dose of the toxic moiety, and (2) the target tissue dose and observed responses (Andersen et al. 

1987; Andersen and Krishnan 1994). These models are biologically and mechanistically based and can 

be used to extrapolate the pharmacokinetic behavior of chemical substances from high to low dose, from 

route to route, between species, and between subpopulations within a species. The biological basis of 

PBPK models results in more meaningful extrapolations than those generated with the more conventional 

use of uncertainty factors. 

The PBPK model for a chemical substance is developed in four interconnected steps: (1) model 

representation, (2) model parametrization, (3) model simulation, and (4) model validation (Krishnan and 

Andersen 1994). In the early 1990s, validated PBPK models were developed for a number of 

toxicologically important chemical substances, both volatile and nonvolatile (Krishnan and Andersen 

1994; Leung 1993). PBPK models for a particular substance require estimates of the chemical substancespecific physicochemical parameters, and species-specific physiological and biological parameters. The 

numerical estimates of these model parameters are incorporated within a set of differential and algebraic 

equations that describe the pharmacokinetic processes. Solving these differential and algebraic equations 

provides the predictions of tissue dose. Computers then provide process simulations based on these 

solutions. 

The structure and mathematical expressions used in PBPK models significantly simplify the true 

complexities of biological systems. If the uptake and disposition of the chemical substance(s) are 

adequately described, however, this simplification is desirable because data are often unavailable for 

many biological processes. A simplified scheme reduces the magnitude of cumulative uncertainty. The 

adequacy of the model is, therefore, of great importance, and model validation is essential to the use of 

PBPK models in risk assessment. 

PBPK models improve the pharmacokinetic extrapolations used in risk assessments that identify the 

maximal (i.e., the safe) levels for human exposure to chemical substances (Andersen and Krishnan 1994). 

PBPK models provide a scientifically sound means to predict the target tissue dose of chemicals in 

humans who are exposed to environmental levels (for example, levels that might occur at hazardous waste 

sites) based on the results of studies where doses were higher or were administered in different species. 

Figure 5-5 shows a conceptualized representation of a PBPK model. 
PBBs and PBDEs 268 

5. HEALTH EFFECTS 

Figure 5-5. Conceptual Representation of a Physiologically Based Pharma 

cokinetic (PBPK) Model for a Hypothetical Chemical Substance 

Source: adapted from Krishnan et al. 1994 

Note: This is a conceptual representation of a physiologically based pharmacokinetic (PBPK) model for a hypothetical 

chemical substance. The chemical substance is shown to be absorbed via the skin, by inhalation, or by ingestion, 

metabolized in the liver, and excreted in the urine or by exhalation. 

Lungs

Liver

Fat

Slowly

perfused

tissues

Richly

perfused

tissues

Kidney

Skin

V

E

N

O

U

S

B

L

O

O

D

A

R

T

E

R

I

A

L

B

L

O

O

D

V max K m

Ingestion

Inhaled chemical Exhaled chemical

G I

Tract

Feces

Urine

 Chemicals in air

contacting skin
PBBs and PBDEs 269 

5. HEALTH EFFECTS 

Polybrominated Biphenyls. A PBPK model that incorporates tissue volume, affinity for PBBs, and rate 

of perfusion was developed to describe the distribution and body burden of the major component of 

FireMaster PBB mixtures, 2,2',4,4',5,5'-hexabromobiphenyl (BB 153), in the rat (Tuey and Matthews 

1980). The modeling methods are an extension of those used to predict the disposition of PCBs 

(Matthews et al. 1977). The model predicts that at equilibrium, changes in the PBB concentration or 

changes in tissue volume of any tissue will lead to a corresponding change in all tissues. For example, if 

the concentration of a PBB congener in the liver is reduced by metabolism or excretion, then the 

concentration of that PBB congener in all tissues will be reduced proportionally. Congeners that cannot 

be readily metabolized (as is the case for BB 153) or excreted will concentrate in adipose tissue, but will 

still circulate to other tissues. Exposure to other tissues will be proportional to the respective tissue/blood 

ratios and the concentration in main storage tissues. This dynamic distribution results in accumulation of 

persistent congeners in all tissues and depletion from all tissues of those congeners that can be cleared. In 

rats orally dosed daily with 14C-2,2',4,4',5,5'-hexabromobiphenyl over a 30-day period, tissue 

concentrations on day 31 were (in increasing order): blood, muscle, liver, skin, and adipose, and were in 

general agreement with those predicted by the PBPK model. When the model was scaled to nonlactating 

humans by adjusting for tissue volume, blood flow, and clearance and rate constant parameters, human 

intake of 9.8 g of the congener from milk consumption over a 230-day period would result in peak 

concentrations of 720 and 2.1 ppm in fat and blood, respectively. The model also predicted that the body 

burden after 5 years would be 5.2 g and the half-life 6.5 would be years. The half-life of 6.5 years 

predicted using the rat-based PBPK model is shorter than mean half-life values of ≈10-15 years estimated 

using human sera data from the Michigan PBB cohort (Blanck et al. 2000b; Lambert et al. 1990; Rosen et 

al. 1995), as discussed in Section 5.4.4 (Elimination and Excretion). The shorter predicted half-life from 

the rat model compared to estimated values based on human sera might be due to differences in adipose 

content between rats and man; fat acts as a depot for these chemicals, and most rat studies use young 

animals with a fat content less than in many people. 

Polybrominated Diphenyl Ethers. No PBPK/PD models were located for PBDEs. 

5.5 MECHANISMS OF ACTION 

5.5.1 Pharmacokinetic Mechanisms 

Polybrominated Biphenyls. The mechanism by which PBBs enter the blood stream from the lungs, skin, 

or gastrointestinal tract is not known and little information is available on how PBBs are distributed in the 
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body. The available data indicate that the absorption mechanism is likely passive diffusion. Results from 

studies of Michigan subjects showed that in the blood stream, 80% of the PBBs were bound to protein 

and 20% was associated with lipids (Greaves et al. 1984). Of the fraction bound to protein, 73% was 

bound to apolipoprotein B and the remaining percent was bound to apolipoprotein A. In an in vitro

model, shown to be representative of environmentally contaminated blood, the distribution of PBBs 

among plasma, erythrocytes, mononucleocytes, and polymorphonucleocytes was 89:9:<1:<1, respectively 

(Roboz et al. 1985). 

In an in vitro study in an adipocyte cell line (3T3L1 cells), >75% of the 2,2',4,4',5,5'-hexabromobiphenyl 

(BB 153) taken up by the cells was associated with subcellular fractions that contained 85% of the 

cellular triglyceride, with only 20% of the compound found in the microsomal plasma-membrane fraction 

(Kraus and Bernstein 1986). This study also found that inhibition of respiration by cyanide at a 

concentration that completely inhibited oxygen consumption did not affect uptake of BB 153, supporting 

the assumption that because of their lipophilic nature, PBBs penetrate membranes by passive diffusion. 

Polybrominated Diphenyl Ethers. No studies were located regarding pharmacokinetic mechanisms for 

PBDEs. 

5.5.2 Mechanisms of Toxicity 

PBBs and PBDEs share some toxicological properties with other structurally similar polyhalogenated 

aromatic compounds, particularly PCBs, chlorinated dibenzo-p-dioxins (CDDs), and chlorinated 

dibenzofurans (CDFs) (Agency for Toxic Substances and Disease Registry 1994, 1998, 2000). However, 

although these chemicals are structurally similar in two dimensions, PBDEs (and PCDEs) differ from the 

other classes on a three dimensional basis. In particular, the oxygen bridge of the ether linkage in the 

diphenyl oxide molecule increases the distance between the biphenyl rings. This apparently reduces 

steric interactions between ortho substituents on the adjacent rings, such that the presence of ortho

bromines is unlikely to present a barrier to rotation that would prevent the two aromatic rings from 

assuming a fully coplanar configuration (Chen et al. 2001; Hardy 2002; Howie et al. 1990). In other 

words, the ether bridge makes PBDEs more non-coplanar in nature, and introducing ortho substitutions 

into PBDEs does not create a spatial impediment for the two phenyl rings to assume a semi-flat position 

with respect to each other, as it does for PBBs or PCBs. Therefore, for PBDEs, the influences of the ether 

bridge and bromine position preclude clearly classifying the congeners as either dioxin-like (coplanar) or 

non-dioxin-like (non-coplanar). This has implications not only for dioxin-type toxicities, which are 
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mediated by the Ah (aryl hydrocarbon) receptor (AhR) pathway, but also for non-dioxin-type effects. For 

example, Chen et al. (2001) found that the induction of CYP1A1 by PBDEs is AhR-mediated, as it is for 

numerous organochlorines, even though PBDEs do not readily adopt the planar conformation usually 

considered characteristic of AhR ligands. Structure-activity relationships have been incompletely 

examined for non-dioxin-like effects of PBDEs such as neurotoxicity. However, based on limited 

available data, it can be speculated that di-ortho-substituted PBDEs might follow the neurotoxic potency 

of ortho-PCBs (Eriksson et al. 2002b; Kodavanti and Derr-Yellin 2002a; Mariussen and Fonnum 2002, 

2003). 

There are also geometrical differences in PCBs, PBBs, and PBDEs due to the higher atomic weight and 

considerably larger molecular volume of bromine compared to chorine (Hardy 2000, 2002). These 

differences contribute to dissimilar physical/chemical properties that can influence the relative 

bioavailability, absorption, tissue accumulation, receptor interactions, and toxicities of the chemicals. For 

example, a comparison of a series of isosteric 3,3',4,4'-tetrahalobiphenyls in rats showed that relative 

toxicity (growth rate and thymic atrophy), AhR binding affinity, and AHH and EROD induction 

potencies increased with increasing bromine substitution (Andres et al. 1983). Possible explanations for 

this effect included the increased polarizability of bromine versus chlorine and differences in the 

electronic, hydrophobic, and hydrogen bonding characteristics of bromine and chlorine (Andres et al. 

1983). The geometrical differences in bromine and chlorine also have implications for understanding the 

mechanism(s) of effects for nondioxin-like PBB congeners, which are not as well characterized as for 

PCBs. In particular, it cannot necessarily be assumed, on the basis of two-dimensional structure, that the 

mechanisms and effects for nondioxin-like PBBs and PCBs are similar. 

Polybrominated Biphenyls. The mechanism of toxicity for PBBs has been extensively studied, but is not 

completely understood (Akoso et al. 1982a, 1982b; Andres et al. 1983; Dannan et al. 1982a, 1982b; 

Goldstein et al. 1979; Parkinson et al. 1983; Render et al. 1982; Robertson et al. 1982; Safe 1984). Many 

PBBs, PCBs, chlorinated dibenzo-p-dioxins (CDDs), chlorinated dibenzofurans (CDFs), and other 

structurally related halogenated aromatic hydrocarbons are believed to share a common mechanism of 

action strongly related to similarities in their structural configuration. Most of what is known regarding 

the mechanism of action of these compounds is based on structure-receptor binding relationships, 

structure-induction relationships, and structure-toxicity relationships (Goldstein and Safe 1989; Safe 

1990). Most of the studies providing this information used parenteral routes of exposure and/or in vitro

test systems, as explained below. It is beyond the scope of this profile to discuss these studies in detail. 
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A limited number of studies have shown that some PBB congeners bind to the cellular AhR, which 

regulates the synthesis of a variety of proteins. This receptor was identified in the cytosol of mouse liver 

cells (Poland et al. 1976) and, subsequently, in extrahepatic tissues of laboratory animals, mammalian cell 

cultures, and human organs and cell cultures (Goldstein and Safe 1989). The structure-binding 

relationships for the coplanar 3,3',4,4',5-pentabromobiphenyl (BB 126), the monoortho substituted 

congener 2,3,3',4,4',5-hexabromobiphenyl (BB 156), and the diortho substituted analog 2,2’,5,5'-tetrabromobiphenyl (BB 52) were examined in rat liver cytosol (Safe et al. 1985). At PBB concentrations 

1,000-fold (10 μM) greater than tetrachlorordibenzo-p-dioxin (TCDD) concentrations (10 nM), the 

coplanar congener completely displaced radiolabeled 2,3,7,8-TCDD from the cytosolic AhR protein, the 

monoortho analog partially displaced the radiolabel, and 2,2',5,5'-tetrabromobiphenyl (BB 52) was the 

least active competitor. The latter congener is relatively nontoxic and does not induce AHH. The Ahbinding characteristics of 3,3',4,4'-tetrabromobiphenyl (BB 77) and 3,3'4,4'5,5'-hexabromobiphenyl, both 

coplanar, were also examined in rat and mice liver cytosol (Millis et al. 1985). The results showed than 

the tetrabromobiphenyl was 10 times more effective than the hexabromobiphenyl in displacing 

radiolabeled 2,3,7,8-TCDD from the receptor. The stereospecific nature of the binding (high affinity seen 

with congeners substituted in both para and two or more meta positions) strongly suggests that a 

biological receptor mediates the responses caused by some PBBs. 

The ability of PBBs to induce hepatic Phase I xenobiotic metabolizing enzymes (cytochrome P-450-

dependent monooxygenases) is well documented (Dannan et al. 1978b, 1982a, 1982b, 1983; Ecobichon et 

al. 1979; Moore et al. 1978, 1979; Parkinson et al. 1983; Robertson et al. 1982; Schramm et al. 1985). 

PBB mixtures were classified as "mixed-type" inducers of hepatic microsomal monooxygenases and 

resembled a mixture of phenobarbital (PB)-like plus 3-methylcholanthrene (MC) as inducers of P-450 

isozymes from CYP1A and CYP2B families. The CYP1A1 and CYP1A2 genes are induced by AhR 

agonists, such as 2,3,7,8-TCDD and MC, and the structure-induction relationships for PBBs as inducers 

of these P-450 isozymes and their related activities have also been determined (Dannan et al. 1983; 

Parkinson et al. 1983; Robertson et al. 1982). For example, when injected intraperitoneally to immature 

male Wistar rats, the coplanar derivatives, 3,4,4'-tribromobiphenyl (BB 37), 3,4,4',5-tetrabromobiphenyl 

(BB 81), 3,3',4,4'-tetrabromobiphenyl (BB 77), 3,3',4,4'5-pentabromobiphenyl (BB 126), and 

3,3',4,4',5,5'-hexabromobiphenyl (BB 169) had a pattern of induction resembling that of MC (Robertson 

et al. 1982). Similar type experiments have shown that monoortho-bromo-substituted analogs of the 

coplanar PBBs, such as 2,3',4,4'-tetrabromobiphenyl (BB 66), 2,3',4,4',5-pentabromobiphenyl (BB 118), 

and 2,3',4,4',5,5'-hexabromobiphenyl (BB 167), exhibit a mixed-type induction activity and resemble 

FireMaster BP-6 in their mode of induction (Dannan et al. 1978b; Parkinson et al. 1983). Yet a third 
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group of PBB congeners, the diortho-bromo analogs of the coplanar PBBs, resemble PB in their mode of 

induction (PB induces the CYP2B1 and CYP2B2 genes). Among the diortho-bromo-substituted PBBs 

studied are 2,2’,5,5'-tetrabromobiphenyl (BB 52), 2,2',4,5,5'-pentabromobiphenyl (BB 101), 

2,2',4,4',5,5'-hexabromobiphenyl (BB 153), and 2,2',3,4,4',5,5'-heptabromobiphenyl (Moore et al. 1979; 

Parkinson et al. 1983). Results from studies with some dibromobiphenyls revealed that 4,4'-dibromobiphenyl (BB 15) resembled PB in its mode of induction (Robertson et al. 1982), whereas 2,2'-dibromobiphenyl (BB 4) had no significant effect on hepatic microsomal drug-metabolizing enzymes (Moore et 

al. 1979). The results of these experiments indicated that coplanar PBB congeners substituted in both 

para and one or more meta positions are MC-type inducers, diortho substituted congeners are PB-type 

inducers, and monoortho analogs of the coplanar PBBs are mixed-type inducers. These results were 

qualitatively similar to those obtained with PCBs and support the idea of a common receptor-mediated 

mechanism of action for PBBs. PBBs are also efficacious inducers of hepatic phase II metabolizing 

enzymes such as glutathione transferases, UDP glucuronyl transferases, and epoxide hydrolase (Parkinson 

et al. 1983; Schramm et al. 1985). For example, when intraperitoneally injected in rats, FireMaster BP-6 

efficaciously induced hepatic glutathione transferases while concomitantly depressing seleniumdependent glutathione perioxidase activity, an important antioxidant enzyme in the liver (Schramm et al. 

1985). 

Many studies that examined structure-induction relationships for several PBB congeners also studied 

structure-toxicity relationships. Thymus and spleen weight were significantly reduced in rats by a series 

of MC-type inducers (Robertson et al. 1982). Further experiments in rats revealed that of a series of PBB 

congeners, only MC-type inducers significantly decreased thymus weight and body weight; PB-type, 

mixed-type, and MC-type inducers increased relative liver weight (Parkinson et al. 1983). Results from 

feeding studies in rats indicate that 3,3',4,4',5,5'-hexabromobiphenyl (BB 169) (MC-type) and 

2,2',4,4',5,5'-hexabromobiphenyl (BB 153) (PB-type) increased liver weight; however, only the MC-type 

inducer decreased body weight and thymic and splenic weight, and caused lymphocytic depletion in the 

thymus (Render et al. 1982). Similar results were obtained when the toxicities of 3,3',4,4'-tetrabromobiphenyl (BB 77) (MC-type) and 2,2'5,5'-tetrabromobiphenyl (weak PB-type) were compared in rats 

(Robertson et al. 1983a). Only BB 77 caused significant reductions in growth rate and thymus size and 

marked depletion of lymphocytes from the thymic cortex. Results from studies with FireMaster BP-6 

revealed that the pattern of toxic responses and the magnitude of the responses attributed to this mixture 

are consistent with it being composed of both MC-type and PB-type congeners; the most toxic responses 

being attributed to the MC-type congeners (Akoso et al. 1982a, 1982b; Dannan et al. 1982a, 1982b; 

Ecobichon et al. 1979; Parkinson et al. 1983; Render et al. 1982). These results suggest a correlation 
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between immunological and hepatic effects and the ability to induce AHH activity and that the most toxic 

congeners are those that resemble the structural configuration of 2,3,7,8-TCDD. This relationship further 

supports the idea of a common receptor-mediated mechanism of action. Other PBB congeners (orthosubstituted) induce other types of effects, such as neurotoxicity, by yet unknown but AhR independent 

mechanisms. 

Some information on structure-promotion relationships for PBBs is available from studies that used twostage liver and skin carcinogenesis models. In the liver promotion studies, development of enzymealtered hepatic foci (putative preneoplastic lesions) was assessed in rats that were partially 

hepatectomized, initiated with diethylnitrosamine and promoted with PBBs (Buchmann et al. 1991; 

Dixon et al. 1988; Evans and Sleight 1989; Jensen and Sleight 1986; Jensen et al. 1982, 1983; Sleight 

1985). Both MC-type (3,3',4,4',5,5'-hexabromobiphenyl [BB 169] and 3,3',4,4'-tetrabromobiphenyl 

[BB 77]) and PB-type (2,2',4,4',5,5'-hexabromobiphenyl [BB 153]) congeners showed hepatic promoting 

activity with varying potencies. FireMaster BP-6 was a more effective promoter than its major 

constituent congener BB 153 (Jensen et al. 1982), which also indicates that other congeners are very 

effective as promoters, or possibly that the combination of congeners with mixed- or PB-type activity 

have a synergistic or additive effect. Although both MC- and PB-type congeners promote two-stage 

hepatic tumor activity, it appears that the MC-type congeners may exert their effects indirectly by causing 

hepatotoxic (cytotoxic effects and necrosis), whereas the PB-type congeners may act as mitogens 

(stimulate cellular growth and division). In a skin tumor assay, HRS/J hairless mice were initiated with 

MNNG and promoted with PBBs (Poland et al. 1982). FireMaster FF-1 and BB 169 were effective skin 

tumor promoters, but BB 153 showed no activity, suggesting that, unlike rat liver tumor promotion, 

promoter activity in the mouse skin tumor model is AhR-dependent. Another indication that promotion 

of tumors by PBBs is not solely an Ah-receptor mediated process is provided by the results of an in vitro

gap junctional intercellular intercommunication assay (Kang et al. 1996). Gap junctional intercellular 

intercommunication in normal human breast epithelial cells was inhibited by 2,2',4,4',5,5'-hexaCB 

(CB 153) in a dose-dependent manner, but not by the coplanar congener 3,3',4,4',5,5'-hexaCB (CB 169). 

Inhibition of gap junctional intercellular communication is generally regarded as a mechanistic marker for 

tumor promotion (as well as several other toxicological end points). 

Expression of the dioxin-type toxic response, which is species and strain dependent, is initiated by the 

binding of individual congeners with the AhR. The responsiveness of a particular organ or cell depends 

on the affinity of the receptor for the ligand molecule (Goldstein and Safe 1989). For example, certain 

inbred strains of mice, typified by C57BL/6J, have a cytosolic AhR protein with a relatively high binding 
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affinity for inducers of benzo[a]pyrene hydroxylase such as 3-MC, ß-naphthoflavone, 2,3,7,8-TCDD, and 

other isostereomers of 2,3,7,8-TCDD. In contrast, other inbred mouse strains, such as DBA/2J, have 

reduced AhR binding affinity. Responsiveness to aromatic hydrocarbons is inherited in a simple 

autosomal dominant mode. Nonresponsiveness has been attributed to a mutation resulting in a receptor 

with a diminished affinity (Okey et al. 1989). This defective receptor is almost completely unresponsive 

to weak inducers such as 3-MC and has reduced sensitivity to more potent inducers such as 

2,3,7,8-TCDD. Studies with PBBs have shown that treatment of C57BL/6J and DBA/2J with FireMaster 

BP-6 resulted in the induction of hepatic microsomal benzo[a]pyrene hydroxylase only in the C57BL/6J 

strain and aminopyrine N-demethylase (PB inducible) in both strains of mice (Robertson et al. 1984a). 

However, 3,3',4,4'-tetrabromobiphenyl (BB 77), a more potent MC-type inducer than the BP-mixture, 

induced benzo[a]pyrene hydroxylase in both strains of mice but did not induce aminopyrine 

N-demethylase in either strain of mice. Also, after treatment with the dioxin-like congener BB 77, thymic 

atrophy was only observed in the responsive strain (Robertson et al. 1984a). In general, studies 

summarized in Section 5.2 in which more than one strain was tested (mice or other species) do not 

address the possible strain-dependency of the toxic responses observed. It must be mentioned, however, 

that differences in the response between tissues, strains, or species, do not exclusively indicate differences 

in receptor affinities, but most likely reflect the fact that the battery of enzyme activities (see below) 

controlled by the Ah locus varies within the tissue, strain, and animal species. 

Initial binding of a PBB congener to the AhR is followed by an activation or transcription step and 

subsequent accumulation of occupied nuclear receptor complexes. These complexes interact with a 

specific DNA sequence in the CYP1A1 gene (which regulates the expression of cytochrome P-450IA1 

isozymes), changing its secondary or supersecondary structure (Elferinck and Whitlock 1990), which 

leads to enhancement of the CYP1A1 gene expression. Newly synthesized enzymes and macromolecules 

resulting from the pleiotropic response to the ligand-receptor complex are believed to be responsible for 

many of the effects caused by PBBs and other halogenated aromatic hydrocarbons. In other words, the 

binding of a congener to the AhR initiates a transcriptional upregulation of a battery of genes that 

modulates biochemical and endocrine pathways, cell cycle regulation (e.g., apoptosis, proliferation, and 

differentiation), morphogenesis, oxidative stress response, and other processes, and is ultimately 

expressed as a diverse spectrum of well characterized toxic responses. 

No studies were located regarding the mechanism of endocrine effects (thyroid toxicity, estrogenicity) of 

PBBs. 
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Polybrominated Diphenyl Ethers. As discussed in the introduction to this section, bromination at the 

ortho position does not appear to significantly change the planarity of PBDE molecules or their biological 

effects. Structure-activity studies have shown that some PBDE congeners can bind to the AhR, although 

binding affinities and induction of AhR-mediated responses are very weak or negligible, particularly for 

commercial PBDE mixtures and environmentally relevant congeners. 

For example, Meerts et al. (1998) indirectly examined the AhR-mediated (dioxin-like) properties of 

17 PBDE congeners in a recombinant H4II rat hepatoma cell line showing AhR mediated expression of a 

luciferase reporter gene. The tested congeners varied from dibromo-substituted to heptabromosubstituted, and with the exception of 4,4'-diBDE (BDE 15)and 3,3',4,4'-tetraBDE (BDE 77), all had at 

least one ortho substitution. Seven of the congeners showed luciferase expression, indicating their ability 

to activate the AhR. The only discernable pattern of receptor activation that appeared to emerge from 

these results was that greater receptor activation was obtained with the penta- and hexaBDEs than with 

tri- and tetraBDEs. Another study also examined the AhR induction potency of PBDE congeners using 

the in vitro luciferase assay with H4IIE-luc recombinant rat hepatoma cells (Villeneuve et al. 2002). Only 

1 of 10 tested congeners, 3,3’,4,4’,5-pentaBDE (BDE 126), induced a significant AhR-mediated gene 

expression response in the H4IIE-luc cells, but the magnitude of induction was only 13% of that caused 

by TCDD. With the exception of 2,3,3’,4,4’-pentaBDE (BDE 105), which induced a response of 1.7% of 

the TCDD maximum, no other congener, including the environmentally prominent congeners 

2,2',4,4'-tetraBDE (BDE 47), 2,2',4,4',5-pentaBDE (BDE 99), and 2,2',4,4',5,5’-hexaBDE (BDE 153), 

yielded a response greater than 1% of TCDD. Overall, the tested PBDE congeners were at least 

200,000 times less potent than TCDD for inducing AhR-mediated gene expression in this test system. 

Chen et al. (2001) studied the affinities of a series of 18 PBDE congeners and 3 commercial PBDE 

mixtures for rat hepatic AhR by using competitive AhR-ligand and EROD induction assays. The analysis 

showed that both the congeners and octa- and pentaBDE commercial mixtures had binding affinities 10-2–

10-5 times that of 2,3,7,8-TCDD. The congener with the highest affinity among the tested congeners was 

2,2',3,4,4'-pentaBDE (BDE 85), although its relative binding affinity was only 2% that of 2,3,7,8-TCDD. 

No binding activity could be determined for the decaBDE mixture. In contrast with PCBs, the binding 

affinities did not appear to relate to the planarity of the molecule, which according to Chen et al. (2001), 

was possibly due to the fact that the large size of bromine atoms expands the receptor binding site. The 

dioxin-like activity of the PBDE congeners and commercial mixtures was subsequently more completely 

characterized, by determining whether they act as AhR agonists or antagonists at sequential stages of the 

AhR signal transduction pathway leading to CYP1A1 in rat hepatocytes (Chen and Bunce 2001). 
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Congeners 3,3’,4,4’-tetraBDE (BDE 77), 2,3’,4,4’,6-pentaBDE (BDE 119), and 3,3’,4,4’,5-pentaBDE 

(BDE 126) were moderately active towards DRE (dioxin response element) binding and induced 

responses of both CYP1A1 mRNA and CYP1A1 protein equivalent to the maximal response of TCDD, 

although at concentrations 3–5 orders of magnitude greater than TCDD. These congeners showed 

additive behavior towards DRE binding with TCDD (i.e., an increased response compared to TCDD 

alone), whereas most of the other congeners antagonized the action of TCDD. Congeners 

2,2’,4,4’,6-pentaBDE (BDE 100), 2,2’,4,4’,5,5’-hexaBDE (BDE 153), and 2,2’,3,4,4’,5’,6-heptaBDE 

(BDE 183) were very weak activators of DRE binding, and other congeners and the three commercial 

BDE mixtures were inactive. In particular, the environmentally prominent congeners 2,2’,4,4’-tetraBDE 

(BDE 47) and 2,2’,4,4’,5-pentaBDE (BDE 99) were among the least active with respect to dioxin-like 

behavior (i.e., were inactive at all stages of signal transduction), and the commercial pentaBDE mixture 

had negligible EROD induction activity. The PBDE congeners that bound most strongly to the AhR were 

also the strongest inducers of CYP1A1 mRNA and CYP1A1 protein, indicating that the induction of 

CYP1A1 was AhR-mediated. Considering all end points evaluated in the Chen et al. (2001) and Chen 

and Bunce (2003) studies, it was concluded that the relative induction potencies (REPs) of the most active 

PBDEs toward CYP1A1 are ≈10-4 that of TCDD (similar to some mono-ortho-PCBs and two orders of 

magnitude less than those of coplanar PCBs), and the REPs for the environmentally prominent congeners 

are essentially zero. 

The enzyme induction properties of PBDEs have been less studied than for other structurally similar 

chemicals. Existing information suggests that PBDEs can be classified as mixed-type inducers of hepatic 

microsomal monooxygenases, although the mixed induction properties of the commercial mixtures are 

likely due to contamination with PBDDs/PBDFs (Darnerud et al. 2001; de Wit 2002; Hardy 2002b). Few 

studies have examined the structure-induction relationships for PBDEs. Chen et al. (2001) examined the 

ability of 12 PBDE congeners and 3 commercial mixtures to induce EROD activity in chick and rat 

hepatocytes, in liver cell lines from rainbow trout, rat, and human, and in a human intestinal cell line. The 

number of bromine substitutions in the congeners tested ranged from 3 to 7. In all cell types, 

3,3',4,4'-tetraBDE (BDE 77), 2,2',4,4',6-pentaBDE, 2,3',4,4'-tetraBDE (BDE 66), and 3,3',4,4',5-pentaBDE (BDE 126) were the strongest inducers. Congeners 2,2',4,4',5,5'-hexaBDE (BDE 153) and 

2,2',3,4,4',5',6-heptaBDE were weak inducers in all cell types, whereas BDE 66 and 2,2',3,4,4'-pentaBDE 

(BDE 85) were very weak inducers in rat hepatocytes and inactive in the other cells. Congeners 

2,2',4,4'-tetraBDE (BDE 47) and 2,2'4,4',5-pentaBDE, which are prominent in the environment, were not 

inducers in any cell line, and neither were 2,4,4'-triBDE (BDE 28), 2,2',4,4',5,6'-hexaBDE, or the penta-, 

octa-, or decaBDE mixtures. For those congeners that had measurable EROD induction activity, their 
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relative potencies were 10-3–10-6 that of 2,3,7,8-TCDD. In general, the EROD induction activity 

paralleled the strength of the AhR binding with the notable exception of 2,2',3,4,4'-pentaBDE (BDE 85), 

which despite its relatively strong AhR binding affinity (see above), showed no evidence of activating the 

AhR to its dioxin-response element (DRE) binding form and was only a weak EROD inducer. 

Relevant information on structure-toxicity relationships is available for PCDEs. Howie et al. (1990) 

examined the immunotoxic potencies of various PCDE congeners on the inhibition of the splenic PFC 

response to SRBC antigen and found the following potency order: 2,3,3',4,4',5-hexaCDE > 

3,3',4,4',5-pentaCDE > 2, 3',4,4',5-pentaCDE > 3,3',4,4'-tetraCDE > 2,2',4,4',5,5'-hexaCDE > 

2,2',4,5,5'-pentaCDE > 2,2',4,4',5,6'-hexaCDE. In general, this potency order paralleled their potencies as 

inducers of hepatic microsomal AHH and EROD. Worth noticing is the fact that the resulting ranking 

order of potency did not follow the order that would have been expected for a response known to be AhRmediated, such as the inhibition of the PFC response to challenge with SRBC antigen. For example, the 

laterally substituted congeners 3,3',4,4'-tetraCDE and 3,3',4,4',5-pentaCDE were less immunotoxic than 

their respective monoortho-substituted analogs; this was true also for their enzyme induction potencies. 

These findings showed that increasing ortho-substitution is less effective in reducing the “dioxin-like” 

activity of these compounds. Howie et al. (1990) suggested that the ether bridge in the polyCDE 

molecules increases the bond length between the two phenyl rings, thus diminishing the effects of ortho

substituents on the biochemical and toxic potencies of these compounds. 

Evidence for thyroid hormone involvement in PBDE toxicity includes observations in rats and mice that 

were orally exposed to commercial mixtures of deca-, octa-, or pentaBDE (see Section 5.2.2.2, Endocrine 

Effects). DecaBDE appears to be much less potent than the lower brominated mixtures and the main 

effects include (1) histological changes in the thyroid indicative of glandular stimulation (e.g., follicular 

cell hyperplasia similar to that induced by a hypothyroid state) (IRDC 1976; Norris et al. 1973, 1975b; 

NTP 1986; WIL Research Laboratories 1984), and (2) decreased serum thyroxine (T4) levels with no 

accompanying changes in serum TSH (Darnerud and Sinjari 1996; Fowles et al. 1994; Hallgren and 

Darnerud 1998;WIL Research Laboratories 1984; Zhou et al. 2001, 2002). Considering these data, the 

structural resemblance of PBDEs to T4, and information from studies of individual congeners as 

summarized below, it is hypothesized that, depending on dose, duration, and mixture/congener, PBDEs 

can disrupt the production, transport, and disposition of thyroid hormones. 

The mechanism(s) by which PBDEs decrease serum T4 levels is is unclear. The apparent lack of effect of 

PBDEs on serum TSH suggests that direct effects on the thyroid leading to inhibition of T4 synthesis are 
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unlikely. PBDEs are hepatic microsomal enzyme inducers, but there is little evidence that increased 

enzyme activity leads to greater clearance of thyroid hormones. The induction of hepatic UDPGT by 

PBDEs has been demonstrated in several studies (Fowles et al. 1994; Hallgren et al. 2001; Zhou et al. 

2001, 2002) and this could increase the UDPGT-catalyzed deactivation and excretion of T4 (i.e., the 

conjugation of T4 with glucuronic acid). An indication that increased UDPGT activity may not be the 

main mechanism for the reduced T4 levels is provided by Hallgren et al. (2001), who found that exposure 

to ≥18 mg/kg/day pentaBDE for 14 days caused serum T4 reductions in both mice and rats with no effect 

on UDPGT activity in the mice, and increased UDPGT in the rats only at higher dose levels. In contrast, 

the decreases in serum T4 correlated with the induction of microsomal phase I enzymes (EROD and 

MROD). As discussed below, increased microsomal enzyme activity could also increase the formation of 

hydroxylated PBDE metabolites that can bind to T4 plasma transport proteins. This would serve to 

increase the number of occupied sites on T4-binding proteins and subsequently result in decreased serum 

levels of T4; however, this mechanism is not fully elucidated. 

The possible interaction of PBDEs with T4 binding to human transthyretin (TTR) was investigated in an 

in vitro competitive binding assay (Meerts et al. 1998, 2000). Testing of 17 congeners, ranging from dito heptaBDEs, showed that none of the parent compounds competed with T4 for binding to human TTR. 

Incubation of the congeners with rat liver microsomes induced by PB (CYP2B enriched), β-napthoflavone (CYPIA enriched), or clofibrate (CYP4A3 enriched) indicated that metabolism is necessary to 

compete with T4-TTR binding and that potency is likely to be both congener and metabolic enzymespecific. The CYP2B-enriched liver microsomes were the most potent, causing 9 of the 17 congeners to 

generate metabolites (not identified) that were effective in displacing T4 from TTR (60% inhibition): 

4,4'-diBDE (BDE 15), 2,4,4'-triBDE (BDE 28), 2,4,6-triBDE (BDE 30), 2,2',4,4’-tetraBDE (BDE 47), 

2,2',4,6'-tetraBDE (BDE 51), 2,4,4',6-tetraBDE (BDE 75), 3,3,4,4'-tetraBDE, 2,2',4,4',6-pentaBDE, and 

2,3',4,4',6-pentaBDE (BDE 119). No T4-TTR inhibition occurred with the higher brominated diphenyl 

ethers (i.e., 2,2',3,4,4',5'-hexaBDE (BDE 138), 2,2',4,4',5,5'-hexaBDE (BDE 153), 2,3,4,4',5,6-hexaBDE 

(BDE 166), and 2,3,3',4,4',5,6-heptaBDE), although it was not verified that these PBDEs were 

metabolized during the in vitro microsomal incubations. Three pure hydroxylated PBDEs, synthesized 

for their structural resemblance with the thyroid hormones 3,5-diiodothyronine (3,5-T2), 3,3',5-triiodothyronine (T3), and 3,3',5,5'-tetraiodothyronine (T4), were also tested in the T4-TTR competition binding 

assay. The relative potencies showed that the T4-like (2,6-dibromo-4-[2,4,6-tribromophenoxy]phenol) 

and T3-like (2-bromo-4-[2,4,6-tribromophenoxy]phenol) hydroxylated PBDEs were 1.42- and 1.22-fold 

more potent, respectively, than T4, and the percentage competition at 500 nM exceeded that of the natural 

ligand; the T2-like hydroxylated PBDE (4-[2,4,6-tribromophenoxy]phenol) showed low affinity for TTR 
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(0.41-fold less potent than T4). Because the PBDEs were able to compete with T4-TTR binding only after 

metabolic conversion by induced rat liver microsomes, and considering that thyroid hormones are also 

hydroxy-halogenated diphenyl ethers. Studies with PCBs and hydroxylated derivatives similarly showed 

that the congener patterns most closely resembling the diiodophenolic ring of thyroxine had the highest 

TTR binding activity (Chauhan et al. 2000; McKinney et al. 1987; Rickenbacher et al. 1986). 

Although the findings discussed above suggest that hydroxylation of PBDEs could be involved in the 

mechanism of thyroid toxicity, there are indications that hydroxylated metabolites might not play a large 

role for commercial penta- and octaBDE products. For example, of the in vitro-generated metabolites 

displacing T4 from TTR (Meerts et al. 1998, 2000), only two (2,2’,4,4’-tetraBDE [BDE 47] and 

2,2’,4,4’,6-pentaBDE [BDE 100]) are known to be present in a commercial product (pentaBDE). An in 

vivo study in rats (Om and Klasson-Wehler 1998) showed that BDE 47 is well absorbed (~95%), but 

poorly metabolized, because (1) only 3% of the amount excreted in feces over a 5-day period was in the 

form of metabolites, (2) the parent molecule was the major compound detected in all tissues analyzed 

(and the only brominated compound detectable in kidney, brain, and adipose tissue), and (3) plasma levels 

were low and mainly due to the parent molecule. Additionally, the three pure hydroxylated PBDEs 

synthesized by Meerts et al. (1998, 2000) for their structural similarity to thyroid hormones are not based 

on congeners known to be present in commercial penta- or octaBDE mixture; this suggests that the 

congeners in these mixtures do not share close structural relationships with thyroid hormones. It is 

relevant to note that high concentrations of hydroxy and methoxy metabolites of PBDEs have been 

detected in fish from the Baltic Sea (Asplund et al. 1999), indicating that these compounds can potentially 

directly impact thyroid function in humans consuming fish. 

Three hydroxylated PBDEs, the 4'-hydroxyl derivatives of 1,3,5-triBDE, 1,3,3',5'-tetraBDE, and 

1,3,3',5,5'-pentaBDE, were tested for affinity to the human thyroid hormone receptor proteins THR-α and 

THR-β in vitro (Marsh et al. 1998). These congeners were tested because they theoretically show the 

highest structural similarity to T4 and T3. None of the hydroxylated derivatives effectively competed with 

the thyroid hormones for binding to either receptor (affinities were 4–>1,000 times less than for T4 and 

T3). Because the tested congeners were the most likely to have affinity for the thyroid hormone receptor, 

it was speculated that other hydroxylated PBDE congeners will have even lower potential for receptor 

binding. DecaBDE (not hydroxylated) had no effect on thyroid hormone receptor-mediated 

transcriptional activation by T3 in HeLaTRDR4-luc human cells; no other congeners were tested in this 

assay (Sakai et al. 2003). 
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The extent that PBDEs affect circulating levels of T4 or T3 is likely to vary with species and rats are 

generally regarded as more sensitive than humans. As discussed in Section 5.5.3, this is thought to be 

related to a smaller and more rapid turnover of the hormone pool in the rat thyroid, and to a more rapid 

clearance of secreted hormone in the rat; that latter being, in part, related to the absence of thyroid binding 

globulin (TBG) in rats (Capen 1997). Transthyretin (TTR) is the major thyroid hormone binding protein 

in rats, whereas TBG is the main binding protein in man and most other mammals. However, although 

TTR is a minor thyroid hormone binding protein in humans, it is the principal protein involved in T4

transport to the brain in both rats and man (Blay et al. 1993; Sinjari et al. 1998). TTR does not transport 

T4 from the bloodstream to the brain, but rather is the main T4 binding protein in cerebrospinal fluid 

(CSF) in both rats and humans. In the rat, T4 is transported to the brain primarily through the blood-brain 

barrier, and not via the choroid plexus and CSF (Blay et al. 1993). 2,2',4,4'-TetraBDE (BDE 47) 

competitively inhibited binding of T4 to sites in rat choroid plexus following in vivo (but not in vitro) 

exposure (Sinjari et al. 1998). Choroid plexus homogenate from rats that were orally treated with 6 or 

18 mg/kg/day for 14 days showed T4 binding that was 80 and 63%, respectively, of that in controls. At 

least one group of mammals is known to exist without TTR (Palha et al. 1997; Schussler 2000). The 

TTR-nul mouse has decreased protein-bound and total T4, normal free T4, and has apparent good health 

(Palha et al. 1997, 2000). Interference with the blood-choroid-plexus-CSF-TTR-mediated route of T4 to 

the brain caused by the absence of TTR did not produce measurable features of hypothyroidism (Palha et 

al. 2000). 

Neurobehavioral developmental alterations have been induced in mice that were neonatally or perinatally 

exposed to individual PBDE congeners, including 2,2',4,4'-tetraBDE (BDE 47), 2,2',4,4',5-pentaBDE 

(BDE 99), 2,2',4,4',5,5’-hexaBDE (BDE 153), and 2,2',3,3’,4,4',5,5’,6,6’-decaBDE (BDE 209) (Branchi 

et al. 2001, 2002; Eriksson et al. 1999, 2001a, 2002a; Viberg et al. 2001a, 2001b, 2002a, 2002b, 2003). 

As detailed in Section 5.2.2.4 Neurological Effects, the main effects were observed at adulthood and 

included reduced spontaneous motor activity, impaired habituation capability, and learning impairment in 

a maze task. The mechanisms for these behavioral and cognitive effects have not been elucidated. One 

possibility involves the well-documented key role of thyroid hormones in brain development. As 

discussed above, in vivo and in vitro exposure to PBDEs has caused changes in hormone levels and other 

thyroid end points in animal models. Some studies suggest that the effects might be related to alterations 

in cholinergic functions. For example, neonatal exposure to a single 8 mg/kg oral dose of BDE 99 on 

PND 10 altered the behavioral response to nicotine, a cholinergic agent, in adult mice (Viberg et al. 

2002b). Neonatal exposure to nicotine and adult exposure to BDE 99 (single 8 mg/kg oral dose at age 

5 months) also affected behavior in mice; the change was not seen in mice only exposed to BDE 99 as 
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adults or mice only exposed to nicotine as neonates (Ankarberg et al. 2001). Adult mice that were 

exposed to single 9 mg/kg oral dose of BDE 153 on PND 10 had a decrease in specific α-Bungarotoxin 

binding sites (cholinergic nicotinic receptors) in the brain hippocampus (Viberg et al. 2001b, 2002a). 

Effects of PBDEs on intracellular signaling processes in rat neuronal cultures and cerebellar fractions 

have been reported. In vitro exposure to commercial pentaBDE mixture DE-71 or tetra congener BDE 47 

stimulated arachidonic acid release in rat cerebellar granule neurons; this effect was not seen with the 

commercial octaBDE product DE-79 (Kodavanti 2003; Kodavanti and Derr-Yellin 2002). The release of 

arachidonic acid appeared to be mediated by the activation of both Ca+2-dependent and Ca+2-independent 

cytosolic phospholipase A2. In vitro exposure to penta mixture DE-71and tetra congener BDE 47 also 

caused translocation of protein kinase C, as indicated by increased phorbol ester binding; octaBDE 

mixture DE-79 did not induce this effect (Kodavanti and Derr-Yellin 2002; Rao et al. 2003). Other 

effects of penta mixture DE-71 and tetra congener BDE 47 included decreases in intracellular calcium 

buffering by microsomes and mitochondria (Kodavanti and Derr-Yellin 2002). The tetra congener BDE 

47 was generally more potent than the DE-71 mixture (mainly comprised of tetra and penta congeners) in 

these tests. Another study found that DE-71 was more toxic than octa and deca congeners in inducing 

cell death and free radical formation in cerebellar granule cells (Reistad et al. 2002). The in vitro uptake 

of the neurotransmitter dopamine into rat brain synaptic vesicles was inhibited by penta mixture DE-71 

mainly tetra and penta congeners), but not by commercial mixtures of octaBDE (DE-79) or decaBDE 

(DE-83R) (Mariussen and Fonnum 2002, 2003). 

The ability of PBDE congeners to induce estrogen receptor (ER)-mediated gene expression in MVLN 

recombinant human breast carcinoma cells, or displace steroid hormones (3

H-testosterone or 3H-estradiol) 

from carp serum proteins, was assessed by Villeneuve et al. (2002). Of 10 tested congeners, including the 

environmentally prominent congeners 2,2',4,4'-tetraBDE (BDE 47), 2,2',4,4',5-pentaBDE (BDE 99), and 

2,2',4,4',5,5’-hexaBDE (BDE 153), induced a significant response in either in vitro assay. Another study 

found that a hydroxylated metabolite of BDE 47, 6-OH-2,2’,4,4’-tetraBDE, did not inhibit aromatase 

(CYP 19) enzyme activity in human H295R adrenocortical cells (Canton et al. 2003). Aromatase is a 

steroidogenic enzyme that mediates the conversion of androgens to estrogens. No other PBDEs were 

tested in this in vitro study. 

The estrogenic and antiestrogenic activities of several PBDE congeners and three hydroxylated PBDEs 

were also assessed in vitro using human breast cell line assays based on ER-dependent luciferase reporter 

gene expression (Meerts et al. 2001). The hydroxylated PBDEs, tribromophenoxy)phenol, 2-bromo-
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5. HEALTH EFFECTS 

4-(2,4,6-tribromophenoxy)phenol, and 2,6-dibromo-4-(2,4,6-tribromophenoxy)phenol, have bromine 

substitution patterns similar to those of the thyroid hormones T2 (3,5-diiodothyronine), T3 (3,3’,5-triiodothyronine), and T4 (3,3’,5,5’-tetraiodothyronine), respectively. Eleven of 17 PBDE congeners showed 

estrogenic activity (dose-dependent luciferase induction) in the ER-CALUX assay with T47D.Luc cells, 

although the most potent PBDE congeners (2,2',4,4',6-pentaBDE [BDE 100] > 2,4,4',6-tetraBDE 

[BDE 75] > 2,2',4,6'-tetraBDE [BDE 51] > 2,4,6-triBDE [BDE 30] > 2,3',4,4',6-pentaBDE [BDE 119]) 

had EC50 values that were 250,000–390,000 times less potent than 17β-estradiol (E2). In contrast, the T3- 

and T2-like hydroxylated PBDEs showed estrogenic potencies exceeding that of E2 (no estrogenic activity 

was induced by the T4-like hydroxylated PBDE). Antiestrogenic potencies were determined in the ERCALUX assay by treating T47D.Luc cells with the PBDEs and hydroxylated PBDEs in the presence of 

E2. Only 2,2',4,4',5,5'-hexaBDE (BDE 153), 2,3,4,4',5,6-hexaBDE (BDE 166), and 2,3,3',4,4',5,6-heptaBDE, which did not induce luciferase activity alone, caused reductions in E2-induced luciferase activity. 

Three of the compounds with potent estrogenic activity in the ER-CALUX assay, 2,4,6-triBDE (BDE 30), 

2,2',4,4',6-pentaBDE (BDE 100), and 4-(2,4,6-tribromophenoxy)phenol, were also tested for estrogenicity 

in ERα-specific and ERβ-specific human embryonic kidney cell lines (293-ERα-Luc and ERβs-Luc cells, 

respectively). The hydroxylated PBDE was potent in the ERα-specific cells (maximum luciferase 

induction similar to E2) and also showed activity in the ERβ-specific cells (maximum 50% induction 

compared to E2), whereas the ERα- and ERβ-specific cell lines were less responsive to 2,4,6-triBDE 

(BDE 30) (34.2 and 7.8% induction compared to E2) and 2,2',4,4',6-pentaBDE (≈20 and <2% relative 

induction). These results indicate that pure and hydroxylated congeners of PBDEs can be agonists of 

both ERα and ERβ receptors and that metabolism of PBDEs may produce more potent pseudoestrogens. 

The common structural features among the estrogenic PBDEs in this study are two ortho (2,6)-bromine 

atoms on one phenyl ring, at least one para-bromine atom (preferably on the same phenyl ring as the 

ortho bromines), and nonbrominated ortho-meta or meta carbons on the other phenyl ring (Meerts et al. 

2001). 

The predominant congeners in biological samples, 2,2’,4,4’-tetraBDE (BDE 47) and 2,2’,4,4’,5-pentaBDE (BDE 99), did not demonstrate estrogenic activity in the Meerts et al. (2001) study. The hydroxyPBDE derivatives that showed minimal activity were para-substituted with a hydroxyl group, whereas the 

para position of the predominant isomers in biological samples is occupied by bromine atom, suggesting 

that the relevance of the results to commercial BDE mixtures is unclear. Additionally, all of the available 

estrogenicity studies were in vitro, and there is, as yet, no evidence either for estrogenicity or antiestrogenicity of PBDEs and/or PBDE metabolites in vivo. 
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5.5.3 Animal-to-Human Extrapolations 

Residue levels of PBBs and PBDEs in humans reflect multiple exposure pathways and congener-specific 

elimination and thus, in general, represent steady state body burdens that do not match the congener 

profiles in the original exposure sources. For example, profiles of PBB and PBDE congeners in human 

milk do not resemble the pattern of any of the commercial mixtures, as illustrated by the finding that the 

major PBDE congener in milk from Swedish mothers was 2,2',4,4'-tetraBDE (BDE 47), which comprised 

approximately 55% of the total PBDEs (Darnerud et al. 1998). As discussed in Chapter 8, residue 

analyses indicate that tetra- to hexa-congeners predominate in humans, aquatic mammals, birds, fish, and 

other biota, indicating that the biological fate of PBB and PBDE congeners is qualitatively similar in 

various animal species. The wildlife residue data also indicate that different species have different tissue 

ratios of congeners, possibly reflective of interspecies differences in metabolic capabilities as well as 

potential differences in exposure. The likelihood of interspecies differences in the quantitative disposition 

of PBBs and PBDEs is illustrated by the observation that metabolism and urinary excretion of a single 

oral dose of BDE 47 was significantly slower in rats than in mice (Orn and Klasson-Wehler 1998). 

Humans are possibly less sensitive than rats to effects of PBDEs on circulating levels of thyroid 

hormones. This difference is thought to derive from the rat thyroid having a smaller store of iodinated 

thyroglobulin that is more easily depleted when the availability of iodide is limited, and from a more rapid 

clearance of T4 from the rat circulation; the latter resulting from rats not having a high affinity binding 

protein for T4 in serum analogous to thyroid-binding globulin (TBG) in humans (Capen 1997). If the 

production of T4 and T3 is impaired sufficiently to deplete the thyroid of stored iodinated thyroglobulin, 

the thyroid cannot produce or secrete amounts of T4 and T3 needed to support physiological demands, 

circulating levels of T4 (free T4) and T3 decrease, and a state of thyroid hormone insufficiency ensues. 

Transthyretin (TTR) is the major thyroid hormone binding protein in rats, but not in man. In most 

mammals, including humans, thyroxin binding globulin (TGB) is the principal thyroid hormone binding 

protein; 74% of the total bound-T4 is bound to TGB, and TTR and albumin bind only 11 and 15%, 

respectively, of the total (Schussler 2000). In contrast to most mammals, the rat utilizes TTR as the major 

T4 plasma binding protein; approximately 75% of T4 in rat serum is bound to TTR and only 25% to 

albumin. Both circulating T3 and T4 are highly protein bound with only a small fraction of their total 

present as free hormone, and this high degree of protein binding serves to maintain equilibrium between 

the extracellular and intracellular pools of these hormones (O’Connor et al. 1999). The increased 

metabolic clearance of serum T4 is thought to involve the induction of hepatic microsomal enzymes. 

Although it is well documented that PBDEs are microsomal enzyme inducers, there is little evidence that 

induction of hepatic enzymes leads to greater clearance of thyroid hormones, indicating that it is 
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5. HEALTH EFFECTS 

misleading to assume that PBDEs are unlikely to affect thyroid function in humans, or that humans are 

less sensitive to these effects than rats. 

Less is known about the relative sensitivities of humans and experimental animals to developmental 

effects of PBDEs. Outstanding uncertainties include potential differences in kinetics of maternal-fetal 

and maternal-infant transfer of PBDEs, as well as potential differences in the degree to which the fetus of 

the human, in comparison to experimental animals, is dependent on maternal thyroid hormone for 

development, particularly during the period of gestation prior to the onset of fetal hormone production. 

5.6 TOXICITIES MEDIATED THROUGH THE NEUROENDOCRINE AXIS 

Recently, attention has focused on the potential hazardous effects of certain chemicals on the endocrine 

system because of the ability of these chemicals to mimic or block endogenous hormones. Chemicals 

with this type of activity are most commonly referred to as endocrine disruptors. However, appropriate 

terminology to describe such effects remains controversial. The terminology endocrine disruptors, 

initially used by Colborn and Clement (1992), was also used in 1996 when Congress mandated the EPA 

to develop a screening program for “...certain substances [which] may have an effect produced by a 

naturally occurring estrogen, or other such endocrine effect[s]...”. To meet this mandate, EPA convened a 

panel called the Endocrine Disruptors Screening and Testing Advisory Committee (EDSTAC), and in 

1998, the EDSTAC completed its deliberations and made recommendations to EPA concerning endocrine 

disruptors. In 1999, the National Academy of Sciences released a report that referred to these same types 

of chemicals as hormonally active agents. The terminology endocrine modulators has also been used to 

convey the fact that effects caused by such chemicals may not necessarily be adverse. Many scientists 

agree that chemicals with the ability to disrupt or modulate the endocrine system are a potential threat to 

the health of humans, aquatic animals, and wildlife. However, others think that endocrine-active 

chemicals do not pose a significant health risk, particularly in view of the fact that hormone mimics exist 

in the natural environment. Examples of natural hormone mimics are the isoflavinoid phytoestrogens 

(Adlercreutz 1995; Livingston 1978; Mayr et al. 1992). These chemicals are derived from plants and are 

similar in structure and action to endogenous estrogen. Although the public health significance and 

descriptive terminology of substances capable of affecting the endocrine system remains controversial, 

scientists agree that these chemicals may affect the synthesis, secretion, transport, binding, action, or 

elimination of natural hormones in the body responsible for maintaining homeostasis, reproduction, 

development, and/or behavior (EPA 1997a). As a result, these chemicals may play a role in altering, for 

example, metabolic, sexual, immune, and neurobehavioral function. Such chemicals are also thought to 
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5. HEALTH EFFECTS 

be involved in inducing breast, testicular, and prostate cancers, as well as endometriosis (Berger 1994; 

Giwercman et al. 1993; Hoel et al. 1992). 

Concern has been raised that many industrial chemicals, including PBBs and PDBEs, are endocrineactive compounds capable of having widespread effects on humans and wildlife (Colborn et al. 1993; 

Crisp et al. 1998; Daston et al. 1997; Safe and Zacharewski 1997). Particular attention has been paid to 

the possibility of these compounds mimicking or antagonizing the action of estrogen. Estrogen influences 

the growth, differentiation, and functioning of many target tissues, including female and male 

reproductive systems, such as mammary gland, uterus, vagina, ovary, testes, epididymis, and prostate. In 

addition, there is evidence that some of these environmentally-persistent chemicals alter the thyroid 

hormone system, which is a very important system for normal structural and functional development of 

sexual organs and the brain. 

Polybrominated Biphenyls. PBBs have the potential to interact with the endocrine system based on 

effects that mainly include changes in levels of thyroid and female reproductive hormones. No studies 

were located that investigated the estrogenic and antiestrogenic activity of PBBs in vitro or in vivo at the 

level of the estrogen receptor. 

The thyroid gland is an unequivocal target of PBBs in animals, and evidence in humans is suggestive of a 

similar relationship. A spectrum of effects has been observed in rats exposed for acute and intermediate 

durations, ranging from decreases in serum levels of T4 and T3 to histological and ultrastructural changes 

in the follicles (Akoso et al. 1982b; Allen-Rowlands et al. 1981; Byrne et al. 1987; Gupta and Moore 

1979; Kasza et al. 1978a; Norris et al. 1975a, 1975b; NTP 1983; Sepkovic and Byrne 1984; Sleight et al. 

1978). The preponderance of these studies tested FireMaster FF-1 or FireMaster BP-6 in rats, although 

chronic exposure to FireMaster FF-1 induced thyroid follicular hyperplasia in mice (NTP 1992). Thyroid 

effects also occurred in offspring of treated rats and pigs (Meserve et al. 1992; Werner and Sleight 1981). 

Effects in workers exposed to unspecified PBBs and/or decabromobiphenyl included increased serum 

FSH, low or borderline low serum T4, and increased thyroid antimicrosomal antibody titers (Bahn et al. 

1980). 

Serum levels of the adrenal hormones corticosterone B, dehydroepiandrosterone, and dehydroepiandrosterone sulfate were decreased in rats fed ≥0.25 mg/kg/day FireMaster BP-6 for 5–7 months 

(Byrne et al. 1988). Serum corticosterone levels and adrenal weight did not change in rats exposed to 

≤6 mg/kg/day of an unspecified PBB mixture for a shorter duration of 20 days (Castracane et al. 1982). 
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5. HEALTH EFFECTS 

Increased menstrual cycle duration and prolonged implantation bleeding were observed in female 

monkeys fed FireMaster FF-1 in approximate daily dose levels of 0.012 mg/kg for 7 months before 

breeding and during pregnancy (Allen et al. 1979; Lambrecht et al. 1978). A corresponding decrease in 

serum levels of progesterone suggests that the reproductive effects in the monkeys are related to PBBinduced endocrine imbalance. Implantation was completely blocked in 40–67% of female rats treated 

with FireMaster BP-6 by gavage in dose levels ≥28.6 mg/kg on alternate days between gestation 

days 0 and 14 (Beaudoin 1979). 

Delayed vaginal opening, an effect suggesting retarded sexual maturation, was observed in F1 generation 

rats whose only PBB exposure was from the mothers fed a diet providing 5 mg/kg/day FireMaster FF-1 

from day 8 of pregnancy until weaning at 28 days postpartum (McCormack et al. 1981). 

Two studies of women exposed during the Michigan contamination episode found no associations 

between serum levels of PBBs and breast feeding (Blanck et al. 2000b; Thomas et al. 2001). 

Determinants of PBB serum decay were investigated in women who had a mean initial PBB level of 

20.9 ppb, a mean time between the first and last measurement of 4.2 years, and at least two measurements 

taken when they were not pregnant (Blanck et al. 2000b). The median PBB half-life was estimated to be 

13.5 years. Subject-specific decay rates were regressed on various predictor variables. Results included 

the finding that breast feeding, as either a continuous variable or as categorized by duration (<3, 3–9, or 

>9 months), was not associated with serum PBB decay, although increasing number of pregnancies was 

weakly associated with a slower rate of serum PBB decay (the effect had borderline statistical 

significance). Additional information on the design and results of this study is provided in Section 5.8.1. 

Thomas et al. (2001) found no relationship between serum levels of PBBs and the frequency and duration 

of lactation in Michigan women. Characteristics of the study cohort included a mean initial serum PBB 

level of 17.5 ppb, an estimated mean serum PBB level at delivery of 9.4 ppb, mean duration of breastfeeding as main source of nutrition of 2.6 months, and mean total duration of breast feeding of 

4.1 months. Exposure was treated as a categorical variable by dividing the women into groups of low 

(reference) exposure (≤1 ppb), moderate exposure (>1–≤7 ppb), and high exposure (>7 ppb), and three 

outcomes of interest were analyzed: (1) the decision to breast feed (yes/no), (2) the duration (months) of 

breast-feeding as the main source of nutrition, and (3) the total duration (months) of breast-feeding. None 

of the three outcomes was significantly associated with serum PBB levels. Additional information on the 

design and results of this study is provided in Section 5.2.2.5. 
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The issue of breast cancer has received attention following reports of high levels of organochlorine 

compounds in breast cancer patients. A relationship between increasing serum levels of PBBs and 

increasing risk of breast cancer was indicated in case-control studies of women exposed during the 

Michigan contamination episode (Henderson et al. 1995; Hoque et al. 1998), but the results are only 

suggestive due to factors such as small number of cases, insufficient information on known breast cancer 

risk factors, and confounding exposures to other organochlorine chemicals. The evidence for an 

association between breast cancer and PCBs is also not conclusive (Agency for Toxic Substances and 

Disease Registry 2000), and the hypothesis that environmental exposure to PCBs can cause breast cancer 

in humans is controversial (Safe and Zacharewski 1997; Wolff and Toniolo 1995). Overall, the evidence 

for an association between breast cancer and PBBs is inconclusive and needs further study. 

Polybrominated Diphenyl Ethers. Results of in vitro estrogen receptor and thyroid hormone transport 

protein binding assays, as well as in vivo thyroid hormone homeostasis studies in animals, suggest that 

there is a potential for some PBDEs to disrupt thyroid and other endocrine system functions in humans. 

The estrogenic and antiestrogenic activities of 17 PBDE congeners and 3 hydroxylated PBDEs were 

tested in vitro using human breast cell line assays based on ER-dependent luciferase reporter gene 

expression (Meerts et al. 2001). Eleven of 17 PBDE congeners showed estrogenic activity (dosedependent luciferase induction) in the ER-CALUX assay with T47D.Luc cells, although the most potent 

PBDE congeners (2,2',4,4',6-pentaBDE > 2,4,4',6-tetraBDE [BDE 75] > 2,2',4,6'-tetraBDE [BDE 51] > 

2,4,6-triBDE [BDE 30] > 2,3',4,4',6-pentaBDE [BDE 119]) had EC50 values that were 250,000–

390,000 times less potent than that of the natural ligand, 17β-estradiol (E2). In contrast, two of the 

hydroxylated PBDEs, 4-(2,4,6-tribromophenoxy)phenol, 2-bromo-4-(2,4,6-tribromophenoxy)phenol 

(which have bromine substitution patterns similar to the thyroid hormones T2 [3,5-diiodothyronine] and 

T3 [3,3’,5-triiodothyronine], respectively) had estrogenic potencies exceeding that of E2. Three of the 

compounds with potent estrogenic activity in the ER-CALUX assay, 2,4,6-triBDE (BDE 30), 2,2',4,4',6-

pentaBDE, and 4-(2,4,6-tribromophenoxy)phenol, were also tested for estrogenicity in ERα-specific and 

ERβ-specific human embryonic kidney cell lines (293-ERα-Luc and ERβs-Luc cells, respectively). The 

hydroxylated PBDE was potent in the ERα-specific cells (maximum luciferase induction similar to E2) 

and also showed activity in the ERβ-specific cells (maximum 50% induction compared to E2), whereas 

the ERα- and ERβ-specific cell lines were less responsive to BDE 30 and 2,2',4,4',6-pentaBDE. These 

results indicate that pure congeners of PBDEs can be agonists of both ERα and ERβ receptors and that 

metabolism to hydroxylated PBDEs may increase estrogenic potency. However, no estrogenic activity 

was demonstrated by the predominant congeners in biological samples, 2,2’,4,4’-tetraBDE (BDE 47) and 
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5. HEALTH EFFECTS 

2,2’,4,4’,5-pentaBDE (BDE 99). The hydroxy-PBDE derivatives that showed minimal activity were 

para-substituted with a hydroxyl group, whereas the para position of the predominant isomers in 

biological samples is occupied by bromine atom, suggesting that the relevance of the results to 

commercial PBDE mixtures is unclear. Additionally, because all of the available estrogenicity data are in 

vitro, there is, as yet, no evidence either for estrogenicity or anti-estrogenicity of PBDEs and/or PBDE 

metabolites in vivo. 

The same 17 PBDE congeners and three hydroxylated PBDEs were also tested for possible interaction 

with T4 binding to human TTR, a plasma transport protein of thyroid hormones, in an in vitro competitive 

binding assay (Meerts et al. 1998, 2000). None of the pure congeners competed with T4 for binding to 

human TTR. Incubation of the congeners with rat liver microsomes induced by PB (CYP2B enriched), 

β-napthoflavone (CYPIA enriched), or clofibrate (CYP4A3 enriched) indicated that 9 of the 17 pure 

congeners generated metabolites (not identified) that were able to displace T4 from TTR: 4,4'-diBDE 

(BDE 15), 2,4,4'-triBDE (BDE 28), 2,4,6-triBDE (BDE 30), 2,2',4,4’-tetraBDE (BDE 47), 2,2',4,6'-tetraBDE (BDE 51), 2,4,4',6-tetraBDE (BDE 75), 3,3,4,4'-tetraBDE, 2,2',4,4',6-pentaBDE, and 

2,3',4,4',6-pentaBDE (BDE 119). Testing of the three known hydroxylated PBDEs, used for their 

structural resemblance with the thyroid hormones 3,5-diiodothyronine (3,5-T2), 3,3',5-triiodothyronine 

(T3), and 3,3',5,5'-tetraiodothyronine (T4) showed that the T4-like (2,6-dibromo-4-[2,4,6-tribromophenoxy]phenol) and T3-like (2-bromo-4-[2,4,6-tribromophenoxy]phenol) hydroxylated PBDEs were 

1.42- and 1.22-fold more potent, respectively, than T4; the T2-like hydroxylated PBDE (4-[2,4,6-tribromophenoxy]phenol) showed low affinity for TTR (0.41-fold less potent than T4). Studies with hydroxylated 

derivatives of PBBs similarly showed that the congener patterns most closely resembling the diiodophenolic ring of thyroxine had the highest TTR binding activity (Chauhan et al. 2000; McKinney et al. 

1987; Rickenbacher et al. 1986). 

Because PBDEs were able to compete with T4-TTR binding only after metabolic conversion, and 

considering that thyroid hormones are also hydroxy-halogenated diphenyl ethers, the results of the Meerts 

et al. (1998, 2000) study suggest that hydroxylation of PBDEs could be involved in the mechanism of 

thyroid toxicity. However, there are indications that hydroxylated metabolites might not play a large role 

for thyroid effects of commercial penta- and octaBDE mixtures. For example, of the in vitro-generated 

metabolites displacing T4 from TTR (Meerts et al. 1998, 2000), only two (2,2’,4,4’-tetraBDE (BDE 47) 

and 2,2’,4,4’,6-pentaBDE) are known to be present in a commercial mixture (pentaBDE). An in vivo

study in rats (Om and Klasson-Wehler 1998) showed that BDE 47 is well absorbed (~95%), poorly 

metabolized (only 3% of the amount excreted in feces was in the form of metabolites), and occurs in 
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5. HEALTH EFFECTS 

blood and other tissues mainly as the parent compound. Additionally, the three pure hydroxylated PBDEs 

synthesized by Meerts et al. (1998, 2000) for their structural similarity to thyroid hormones are not based 

on congeners known to be present in commercial penta- or octaBDE mixtures; this suggests that the 

congeners in these mixtures do not share close structural relationships with thyroid hormones. In another 

study, 4'-hydroxyl derivatives of 1,3,5-triBDE, 1,3,3',5'-tetraBDE, and 1,3,3',5,5'-pentaBDE were tested 

for affinity to the human thyroid hormone receptor proteins THR-α and THR-β in vitro (Marsh et al. 

1998). These congeners were tested because they theoretically show the highest structural similarity to T4

and T3. None of the hydroxylated derivatives effectively competed with the thyroid hormones for binding 

to either receptor (affinities were 4–>1,000 times less than for T4 and T3). Because the tested congeners 

were the most likely to have affinity for the thyroid hormone receptor, it is likely that other hydroxylated 

PBDE congeners will have even lower potential for receptor binding. 

Studies of serum hormone levels and organ histology indicate that the thyroid is a sensitive target for 

some PBDEs. Reduced serum T4 levels and follicular cell hyperplasia have been consistently observed in 

rats and mice orally exposed to PBDEs. Acute-duration studies showed decreases in serum T4 in rats 

exposed to ≥10 mg/kg/day octaBDE or ≥30 mg/kg/day pentaBDE for 4 days and in rats and mice exposed 

to ≥18 mg/kg/day pentaBDE for 14 days (Darnerud and Sinjari 1996; Hallgren et al. 2001; Zhou et al. 

2001). Effects observed in intermediate-duration studies include thyroid hyperplasia in rats exposed to 

≥8 mg/kg/day octaBDE for 30 days (Norris et al. 1973, 1975a, 1975b) and reduced serum T4 in rats 

exposed to ≥10 mg/kg/day pentaBDE for 90 days (WIL Research Laboratories 1984). Exposure to 

pentaBDE on gestation day 6 through postnatal day 21 caused serum T4 reductions at 30 mg/kg/day in 

maternal rats and at ≥10 mg/kg/day in their offspring (Zhou et al. 2002). Intermediate-duration exposure 

to a 77% decaBDE/22% nonaBDE commercial mixture caused thyroid hyperplasia in rats at doses of 

≥80 mg/kg/day for 30 days (Norris et al. 1973, 1975a, 1975b). Chronic (103-week) exposure to highpurity commercial decaBDE (≥97%) did not induce thyroid histopathological changes in rats at 

≤2,550 mg/kg/day, although follicular cell hyperplasia developed in mice exposed to 2,240 mg/kg/day 

(NTP 1986). 

As summarized above, evidence for thyroid hormone involvement in PBDE toxicity includes 

observations in rats and mice that were orally exposed to commercial mixtures of deca-, octa- or 

pentaBDE. The main effects include (1) histological changes in the thyroid indicative of glandular 

stimulation (e.g., follicular cell hyperplasia similar to that induced by a hypothyroid state) (IRDC 1976; 

Norris et al. 1973, 1975b; NTP 1986; WIL Research Laboratories 1984), and (2) decreased serum 

thyroxine (T4) levels with no accompanying changes in serum TSH (Darnerud and Sinjari 1996; Fowles 
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5. HEALTH EFFECTS 

et al. 1994; Hallgren and Darnerud 1998;WIL Research Laboratories 1984; Zhou et al. 2001, 2002), and 

decaBDE appears to be much less potent than the lower brominated mixtures. Considering these data, the 

structural resemblance of some PBDEs to T4, and information from binding studies of individual 

congeners as summarized above, it is hypothesized that, depending on dose, duration, and 

mixture/congener, PBDEs can disrupt the production, transport, and disposition of thyroid hormones. 

The extent that PBDEs affect circulating levels of T4 or T3 is likely to vary with species, and rats are 

generally regarded as more sensitive than humans. This is thought to be related to a smaller and more 

rapid turnover of the hormone pool in the rat thyroid, and to a more rapid clearance of secreted hormone 

in the rat; that latter being, in part, related to the absence of thyroid binding globulin (TBG) in rats (Capen 

1997). Transthyretin (TTR) is the major thyroid hormone binding protein in rats, whereas TBG is the 

main binding protein in man and most other mammals. However, although TTR is a minor thyroid 

hormone binding protein in humans, it is the principal protein involved in T4 transport to the brain in both 

rats and man (Blay et al. 1993; Sinjari et al. 1998). TTR does not transport T4 from the bloodstream to 

the brain, but rather is the main T4 binding protein in cerebrospinal fluid (CSF) in rats and humans. In the 

rat, T4 is transported to the brain primarily through the blood-brain barrier, and not via the choroid plexus 

and CSF (Blay et al. 1993). 2,2',4,4'-TetraBDE (BDE 47) competitively inhibited binding of T4 to sites in 

rat choroid plexus homogenates following in vivo exposure (Sinjari et al. 1998). 

5.7 CHILDREN’S SUSCEPTIBILITY 

This section discusses potential health effects from exposures during the period from conception to 

maturity at 18 years of age in humans, when all biological systems will have fully developed. Potential 

effects on offspring resulting from exposures of parental germ cells are considered, as well as any indirect 

effects on the fetus and neonate resulting from maternal exposure during gestation and lactation. 

Relevant animal and in vitro models are also discussed. 

Children are not small adults. They differ from adults in their exposures and may differ in their 

susceptibility to hazardous chemicals. Children’s unique physiology and behavior can influence the 

extent of their exposure. Exposures of children are discussed in Section 8.6 Exposures of Children. 

Children sometimes differ from adults in their susceptibility to hazardous chemicals, but whether there is 

a difference depends on the chemical (Guzelian et al. 1992; NRC 1993). Children may be more or less 

susceptible than adults to health effects, and the relationship may change with developmental age 
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5. HEALTH EFFECTS 

(Guzelian et al. 1992; NRC 1993). Vulnerability often depends on developmental stage. There are 

critical periods of structural and functional development during both prenatal and postnatal life and a 

particular structure or function will be most sensitive to disruption during its critical period(s). Damage 

may not be evident until a later stage of development. There are often differences in pharmacokinetics 

and metabolism between children and adults. For example, absorption may be different in neonates 

because of the immaturity of their gastrointestinal tract and their larger skin surface area in proportion to 

body weight (Morselli et al. 1980; NRC 1993); the gastrointestinal absorption of lead is greatest in infants 

and young children (Ziegler et al. 1978). Distribution of xenobiotics may be different; for example, 

infants have a larger proportion of their bodies as extracellular water and their brains and livers are 

proportionately larger (Altman and Dittmer 1974; Fomon 1966; Fomon et al. 1982; Owen and Brozek 

1966; Widdowson and Dickerson 1964). The infant also has an immature blood-brain barrier (Adinolfi 

1985; Johanson 1980) and probably an immature blood-testis barrier (Setchell and Waites 1975). Many 

xenobiotic metabolizing enzymes have distinctive developmental patterns. At various stages of growth 

and development, levels of particular enzymes may be higher or lower than those of adults, and 

sometimes unique enzymes may exist at particular developmental stages (Komori et al. 1990; Leeder and 

Kearns 1997; NRC 1993; Vieira et al. 1996). Whether differences in xenobiotic metabolism make the 

child more or less susceptible also depends on whether the relevant enzymes are involved in activation of 

the parent compound to its toxic form or in detoxification. There may also be differences in excretion, 

particularly in newborns who all have a low glomerular filtration rate and have not developed efficient 

tubular secretion and resorption capacities (Altman and Dittmer 1974; NRC 1993; West et al. 1948). 

Children and adults may differ in their capacity to repair damage from chemical insults. Children also 

have a longer remaining lifetime in which to express damage from chemicals; this potential is particularly 

relevant to cancer. 

Certain characteristics of the developing human may increase exposure or susceptibility, whereas others 

may decrease susceptibility to the same chemical. For example, although infants breathe more air per 

kilogram of body weight than adults breathe, this difference might be somewhat counterbalanced by their 

alveoli being less developed, which results in a disproportionately smaller surface area for alveolar 

absorption (NRC 1993). 

Children are exposed to PBBs and PBDEs in the same manner as the general population, primarily via 

consumption of contaminated foods. Exposure also may occur by transfer of PBBs and PBDEs that have 

accumulated in women’s bodies to the fetus across the placenta. Because PBBs and PBDEs are lipophilic 

substances, they can additionally accumulate in breast milk and be transferred to nursing infants. 
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Placental transfer, although it may be limited in absolute amounts, is a concern because of possible effects 

of PBBs and PBDEs on sensitive immature tissues, organs, and systems, with potentially serious longlasting consequences. Transfer of PBBs and PBDEs via breast milk could be relatively considerable and, 

like prenatal exposure, has the potential to contribute to altered development. 

Although embryos, fetuses, and nursing infants may be exposed to relatively high amounts of PBBs and 

PBDEs during sensitive periods of development, it is not known if the susceptibility of children to the 

health effects of these chemicals differs from that of adults. Younger children may be particularly 

vulnerable to PBBs and PBDEs because, compared to adults, they are growing more rapidly and are 

generally expected to have lower and distinct profiles of biotransformation enzymes, as well as much 

smaller fat depots for sequestering these lipophilic chemicals. No specific information was located 

regarding the pharmacokinetics of PBBs in children or nutritional factors that may influence the 

absorption of PBBs. 

No biomarkers of exposure or effects for PBBs or PBDEs have been validated in children or in adults 

exposed as children. There also are no biomarkers in adults that identify previous childhood exposure. 

No studies were located regarding interactions of PBBs or PBDEs with other chemicals in children or 

adults. No information was located regarding pediatric-specific methods for reducing peak absorption 

following exposure to PBBs or PBDEs, reducing body burden, or interfering with the mechanism of 

action for toxic effects. In addition, no data were located regarding whether methods for reducing toxic 

effects in adults might be contraindicated in children. 

Polybrominated Biphenyls. Information on health effects of PBBs in children is available from several 

studies of the Michigan contamination episode. A 1976 medical history questionnaire study of 

342 Michigan children likely to have been exposed to PBBs found that the number of subjectively 

reported symptoms of ill health, including several symptoms of neurological effects, did not increase with 

increasing serum PBB levels (assayed in 1976), but rather decreased; general neurological examinations 

did not reveal a pattern of abnormality among the Michigan children (Barr 1980). Studies of fetal 

mortality rates in Michigan (Humble and Speizer 1984) and of physical and neuropsychological 

development in Michigan children exposed during the contamination episode (Schwartz and Rae 1983; 

Seagull 1983; Weil et al. 1981) did not conclusively correlate the ingestion of PBBs with effects, as 

summarized below. 
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Examination of approximately 100 children presumably exposed in utero or in early infancy during the 

peak of the Michigan contamination episode and whose families lived on farms known to be 

contaminated with PBBs has not revealed any consistent or marked abnormalities. No significant 

abnormalities were found by physical and neuropsychological examination of 33 of these exposed 

children when they had a mean age of 37.2 months, compared with a group of 20 age-matched, nonexposed control children (Weil et al. 1981). However, subjective interviews with parents suggested that 

more exposed children than control children had frequent upper respiratory illnesses such as colds, runny 

noses, and sore throats (Weil et al. 1981). Administration of 5 of 18 possible neuropsychological 

development tests from the McCarthy Scales of Children's Abilities to 19 of these exposed children at 

≈2.5–4 years of age showed a statistically significant negative correlation between PBB levels in fat tissue 

and developmental abilities in four of the five tests (Seagull 1983). Subsequent administration of the full 

battery of 18 neuropsychological tests, as well as I.Q. tests, to the same group of children when ≈4–

6 years old, found that the exposed children's performances were within the normal range in all areas 

assessed (Schwartz and Rae 1983). Due mainly to the small data set and the inconsistency of the results, 

the available data do not conclusively establish or eliminate the possibility that in utero and early infancy 

exposure to PBBs might adversely affect the development of human children. 

Neurobehavioral alterations have been observed in animals following gestational and lactational exposure 

to PBBs. Performance deficits in tests of operant behavior were seen in 6-month-old offspring of rats that 

were exposed to ≥0.2 mg/kg/day of FireMaster BP-6 by gavage from day 6 of gestation until weaning 

(Henck and Rech 1986), but not in 75-day-old offspring of rats exposed to ≥0.5 mg/kg/day for 4 weeks 

prior to mating (Geller et al. 1985). Effects on acquisition of forward locomotion, cliff avoidance, cage 

emergence, and open-field activity were found in offspring of rats exposed to ≥0.2 mg/kg/day of 

FireMaster BP-6 in the diet from day 6 of gestation through PND 24 and tested through PND 60 (Henck 

et al. 1994). Testing of mouse offspring at 30–120 days of age following maternal exposure to 

FireMaster FF-1 by gavage on every other day during gestation and through weaning showed altered 

negative geotaxis and avoidance response latencies at ≥3 mg/kg/day and reduced acoustic startle 

responsiveness and motor activity at 10 mg/kg/day (Tilson 1992). 

Animal studies of FireMaster FF-1 and FireMaster BP-6 have also shown that hexabromobiphenyl PBB 

mixtures can induce non-neurological developmental toxicity. Embryolethal effects or increased 

mortality among nursing young were observed in rats (Beaudoin 1977, 1979; Groce and Kimbrough 

1984) and mice (Luster et al. 1980) after oral exposure during gestation and in monkeys after exposure 

before conception and during pregnancy (Allen et al. 1979; Lambrecht et al. 1978). Structural 
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5. HEALTH EFFECTS 

malformations in fetuses, including cleft palate, were also observed in rats (Beaudoin 1977) and mice 

(Corbett et al. 1975) after exposure to these PBBs during gestation. Increased incidences of fetuses with 

extra ribs were found in a study of rats orally exposed to a commercial octabromobiphenyl mixture during 

gestation (Waritz et al. 1977), although decabromobiphenyl was not embryotoxic, fetotoxic, or 

teratogenic in rats (Millischer et al. 1980). Other studies with FireMaster FF-1 and FireMaster BP-6 

found that body weight gain was reduced in the offspring of rats and mice after exposure during gestation 

and/or lactation (Corbett et al. 1975; Groce and Kimbrough 1984; McCormack et al. 1981, 1982c; 

Meserve et al. 1992). Liver effects, including increased liver weight and hepatic cytochrome P-450 

enzymic activity, hepatocyte enlargement, vacuolization, and/or other degenerative changes, occurred in 

the offspring of rats, mice, and swine fed FireMaster FF-1 or FireMaster BP-6 during gestation and/or 

lactation (Chhabra et al. 1993; Moore et al. 1978; NTP 1992; Werner and Sleight 1981). 

Other effects in offspring of animals exposed to PBBs during gestation and lactation include altered 

thyroid hormone levels. Serum T4 levels were reduced in 15-day-old offspring of rats that were exposed 

to 2.5 mg/kg/day FireMaster BP-6 in the diet from GD 0 through PND 15 (Meserve et al. 1992). The 

pups had received pituitary stimulation by an injection of corticotropin-releasing factor or adrenal 

stimulation by an injection of adrenocorticotropic hormone. Serum concentrations of T3 and T4 were 

significantly reduced in 4-week-old nursing offspring of swine that were fed ≥1.25 mg/kg/day dietary 

doses of FireMaster BP-6 during the second half of gestation and throughout lactation (Werner and 

Sleight 1981). These effects in offspring are consistent with evidence that the thyroid gland is an 

unequivocal target of PBBs in adult animals. A spectrum of thyroid effects, ranging from decreases in 

serum T4 and T3 levels to histological and ultrastructural changes in the follicles, has been documented in 

adult rats orally exposed to PBBs (mainly FireMaster BP-6 and FF-1) for acute and intermediate 

durations (Akoso et al. 1982b; Allen-Rowlands et al. 1981; Byrne et al. 1987; Gupta and Moore 1979; 

Kasza et al. 1978a; Norris et al. 1975a, 1975b; NTP 1983; Sepkovic and Byrne 1984; Sleight et al. 1978). 

Additionally, there is suggestive limited evidence of thyroid effects in adult humans; effects in workers 

exposed to unspecified PBBs and/or decaBDE included increased serum FSH, low or borderline low 

serum T4, and increased thyroid antimicrosomal antibody titers (Bahn et al. 1980). 

No information was located on possible immunological effects of PBBs in children, and data in adult 

humans are limited and largely inconclusive. Altered lymphocyte transformation responses among 

populations exposed to PBBs during the Michigan contamination episode have been reported in some 

studies (Bekesi et al. 1978; Roboz et al. 1985), but other investigations were not able to confirm these 

findings (Landrigan et al. 1979; Silva et al. 1979; Stross et al. 1981). No correlation can be established 
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5. HEALTH EFFECTS 

between altered immune parameters and serum PBB levels based on the available data. Exposure to 

PBBs caused altered immune responses in a variety of animal species, which suggests that humans may 

also be affected. Studies in animals, mostly intermediate-duration studies in rodents, showed that a 

variety of immunological parameters such as spleen and thymus weights (Gupta and Moore 1979; Gupta 

et al. 1981; NTP 1983), antibody production (Loose et al. 1981), and lymphoproliferative responses 

(Howard et al. 1980; Luster et al. 1978, 1980) can be affected by treatment with commercial PBB 

mixtures, although some of these effects were only seen at PBB levels that cause overt toxicity (Luster et 

al. 1978, 1980). 

Levels of PBBs in breast milk have been measured in women exposed as a result of the Michigan 

contamination episode. The milk concentrations of PBBs in women from the lower peninsula of 

Michigan (exposed area) were generally higher than in breast milk of females from the upper peninsula 

(farthest from the sources) (Brilliant et al. 1978). PBB levels in breast milk of five women from exposed 

farms ranged from 0.21–92.7 ppm (Cordle et al. 1978; Humphrey and Hayner 1976). On a lipid basis, the 

ratio of PBBs in breast milk to maternal serum was 107–122 to 1 and in adipose tissue to breast milk was 

1.1–1.5 to 1 in a cohort of Michigan residents (Eyster et al. 1983; Landrigan et al. 1979). No monitoring 

information was located on PBBs in breast milk for U.S. populations outside of Michigan. 

Determinants of PBB serum decay were investigated in Michigan women who had a mean initial PBB 

level of 20.9 ppb, a mean time between the first and last measurement of 4.2 years, and at least two 

measurements taken when they were not pregnant (Blanck et al. 2000b). Assuming that PBBs reached 

equilibrium in the body before substantial amounts were eliminated and before the first serum 

measurements were taken, the authors estimated the median PBB half-life to be 13.5 years. Subjectspecific decay rates were regressed on various predictor variables. Results of the analysis included the 

finding that an increasing number of pregnancies between the first and last measurement was associated 

with a slower rate of serum PBB decay (the effect had borderline statistical significance). Breast feeding 

as either a continuous variable or as categorized by duration (<3, 3–9, or >9 months) was not associated 

with serum PBB decay. Additional information on the design and results of this study is provided in 

Section 5.8.1. Another study of women exposed to PBBs during the Michigan contamination episode 

similarly found no relationship between serum levels of PBBs and the frequency and duration of lactation 

(Thomas et al. 2001). Characteristics of the study cohort included a mean initial serum PBB level of 

17.5 ppb, an estimated mean serum PBB level at delivery of 9.4 ppb, a mean duration of breast-feeding as 

the main source of nutrition for 2.6 months, and a mean total duration of breast-feeding of 4.1 months. 

Exposure was treated as a categorical variable by dividing the women into groups of low (reference) 
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5. HEALTH EFFECTS 

exposure (≤1 ppb), moderate exposure (>1–≤7 ppb), and high exposure (>7 ppb). Three outcomes of 

interest were analyzed: (1) the decision to breast feed (yes/no), (2) the duration (months) of breast-feeding 

as the main source of nutrition, and (3) the total duration (months) of breast-feeding. None of the three 

outcomes was significantly associated with serum PBB levels, even after controlling for various 

confounding determinants, including histories of previous breast-feeding and thyroid disorders. 

Additional information on the design and results of this study is provided in Section 5.2.2.5. 

Polybrominated Diphenyl Ethers. No specific information was located on health effects of PBDEs in 

newborn or older children. Thyroid and neurobehavioral alterations have been observed in developing 

animals following pre- and/or postnatal exposure to lower brominated BDEs, indicating that infants and 

children might be particularly susceptible to these effects. The potential for thyroid and 

neurodevelopmental effects of decaBDE is expected to be relatively low due to differences in the 

toxicokinetics and toxicity of decaBDE compared to lower brominated BDEs. 

Thyroid hormones regulate cell proliferation, migration, and differentiation during development, and 

maintenance of normal levels is essential to normal growth and development. Disruption of circulating 

hormone levels can have markedly different effects, depending on the stage of development, and even 

transient disruptions can produce permanent effects. Effects can include mental retardation, impaired 

motor skills, and hearing and speech impediments (Boyages 2000; Fisher and Brown 2000). Several 

factors might contribute to a high vulnerability of the fetus and neonate to lower brominated PBDEs. 

Relatively brief periods of thyroid hormone insufficiency (e.g., 14 days) can produce measurable 

neurological deficits in newborn infants (van Vliet 1999). Furthermore, unlike the adult thyroid gland, 

which contains a relatively large store of T4 that is sufficient to support circulating levels of hormone for 

several months, the neonatal thyroid contains only enough hormone to support circulating levels of 

hormone for ≥1 day (van den Hove et al. 1999; Vulsma et al. 1989). Thus, even acute exposures to a dose 

of lower brominated PBDEs sufficient to suppress thyroid hormone production could potentially result in 

thyroid insufficiency in the neonate. The absorbed dose of lower brominated PBDEs per unit of body 

mass is also likely to be higher in infants compared to adults exposed to similar levels of PBDEs because 

of higher intakes per unit of body mass and exposure from breast milk. It should be noted that screening 

of all newborn children for hypothyroidism is already a widely accepted and legislatively mandated 

practice (LaFranchi 1999; Landenson et al. 2000); newborns are tested for thyroid hormone levels within 

the first few days of life in the United States and most other developed countries; and treatment is started 

immediately if indicated (LaFranchi 1999; Landenson et al. 2000). 
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Although there is no evidence that exposure to PBDEs causes thyroid effects in humans, the thyroid is a 

well-documented target of lower brominated commercial mixtures and single congeners in adult and 

neonatal rats and mice. The main effects in animals are reduced serum T4 hormone levels and follicular 

cell hyperplasia, with no accompanying changes in serum T3 or TSH levels. In adult animals, acuteduration oral exposure caused decreases in serum T4 in rats exposed to ≥10 mg/kg/day octaBDE or 

≥30 mg/kg/day pentaBDE for 4 days, and in rats and mice exposed to ≥18 mg/kg/day pentaBDE for 

14 days (Darnerud and Sinjari 1996; Hallgren and Darnerud 1998; Hallgren et al. 2001; Zhou et al. 2001). 

Effects observed in intermediate-duration oral studies included thyroid hyperplasia in rats exposed to 

≥8 mg/kg/day octaBDE for 30 days (Norris et al. 1973, 1975a, 1975b) and reduced serum T4 in rats 

exposed to ≥10 mg/kg/day pentaBDE for 90 days (WIL Research Laboratories 1984). 

Other animal studies have shown reduced serum levels of thyroid hormones in offspring of rats exposed 

to lower brominated PBDEs during gestation and lactation, as well as in rats orally exposed as weanlings. 

Exposure to pentaBDE from GD 6 through the end of lactation caused serum T4 decreases in maternal 

rats (on GD 20 and PND 22) at 30 mg/kg/day and in their offspring (on PNDs 4 and 14) at ≥10 mg/kg/day 

(Zhou et al. 2002). Assessment of weanling (28-day-old) rats that were orally exposed to PBDEs for 

4 days and evaluated for thyroid hormone changes on the day after the last exposure showed that 

octaBDE caused significantly reduced serum T4 and T3 levels at ≥10 and ≥60 mg/kg/day, respectively 

(Zhou et al. 2001). Similar exposure to pentaBDE caused serum T4 and T3 decreases at higher dose levels 

of ≥30 and ≥100 mg/kg/day, respectively. The results of these studies are consistent with findings of 

reduced serum T4 hormone levels and follicular cell hyperplasia in adult rats and mice confirming that the 

thyroid is a sensitive target of octaBDE and pentaBDE at doses as low as 10 mg/kg/day (Darnerud and 

Sinjari 1996; Hallgren et al. 2001; WIL Research Laboratories 1984). 

Although alterations in thyroid hormone homeostasis can cause neurodevelopmental effects, little specific 

information is currently available on the potential neurotoxic effects of PBDEs. Data are mainly limited 

to the results of three behavioral tests in animals showing some alterations in spontaneous locomotion 

behavior and learning and memory ability in mice that were tested at 2 months of age and as adults 

(4 months), following neonatal exposure (PNDs 3, 10, or 19) to single low oral doses of the congeners 

2,2',4,4'-tetraBDE (BDE 47) or 2,2',4,4',5-pentaBDE (BDE 99) (Eriksson et al. 1998, 1999, 2001, 2002a). 

Effects on spontaneous activity were observed in adult mice treated at either 3 or 10 days of age, but not 

at 19 days of age, suggesting that there is a critical phase of neonatal brain development for the induction 

of behavioral disturbances (Eriksson et al. 1999, 2002a). No studies were located that sufficiently 

evaluated neurological effects of PBDEs in animals exposed as adults. None of the commercial decaBDE, 
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5. HEALTH EFFECTS 

octaBDE, and pentaBDE products have been screened for neurotoxicity using comprehensive test 

batteries that typically include functional observational, motor activity, and neuropathology evaluations. 

There were no indications of neurotoxicity in rats and mice fed high dietary dose levels of a 94–97% pure 

commercial decaBDE mixture for 14 days (≤19,000 mg/kg/day), 13 weeks (≤9,500 mg/kg/day), or 

103 weeks (≤7,780 mg/kg/day), as assessed by overt clinical signs (all exposures) and nervous system 

histopathology (chronic exposure only) (NTP 1986). Although the high doses and extended exposure 

durations in the NTP (1986) studies provided opportunities for the development of effects, neurotoxicity 

is only incompletely evaluated due to the lack of testing for subtle behavioral and other sensitive 

neurological end points. 

The human relevance of the thyroid and neurodevelopmental effects of lower brominated BDEs in 

animals is unclear. Humans are generally regarded as being less sensitive than rats to effects of PBDEs 

on circulating thyroid hormones. This is thought to be related to a smaller and more rapid turnover of the 

hormone pool in the rat thyroid, and to a more rapid clearance of secreted hormone in the rat; the latter 

being, in part, related to the absence of thyroid binding globulin (TBG) in rats (Capen 1997). 

Transthyretin (TTR) is the major thyroid hormone binding protein in rats, whereas TBG is the main 

binding protein in man and most other mammals. However, although TTR is a minor thyroid hormone 

binding protein in humans, it is the principal protein involved in T4 transport to the brain in both rats and 

man (Blay et al. 1993; Sinjari et al. 1998). TTR does not transport T4 from the bloodstream to the brain, 

but rather is the main T4 binding protein in cerebrospinal fluid (CSF) in rats and humans. In the rat, T4 is 

transported to the brain primarily through the blood-brain barrier, and not via the choroid plexus and CSF 

(Blay et al. 1993). Also, mechanism by which lower brominated BDEs cause decreased serum T4 might 

involve hepatic microsomal enzyme induction and consequent increased metabolic formation of hydroxymetabolites, but humans are not particularly sensitive to this effect. Additionally, the animal studies in 

which neurobehavioral alterations were observed were performed using an atypical design that has not 

been validated, and used exposure levels much higher than those likely to be experienced by humans. 

The potential for induction of thyroid and neurodevelopmental effects by decaBDE is low in comparison 

to commercial penta- and octaBDE mixtures due to toxicokinetic and toxicity differences. The lower 

brominated BDEs preferentially accumulate in the body due to their partitioning and retention in lipidrich tissues and lower rates of metabolism and elimination relative to decaBDE (Hardy 2002b; Morck et 

al. 2001, 2003). These characteristics seem to be a function of the number and location of bromines on 

the diphenyl oxide molecule . The tetra- and pentaBDEs appear to be relatively well absorbed, whereas 

the fully brominated decaBDE, a large poorly soluble molecule, is very poorly absorbed (≈1% or less of 
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an oral dose) and rapidly eliminated (≈99% of the dose within 72 hours). Further, decaBDE is 

significantly less toxic than lower brominated BDE mixtures. For example, a study in weanling rats 

showed that acute oral exposure to commercial decaBDE had no effect on thyroid hormones, while 

similar exposures to commercial octaBDE or pentaBDE mixtures caused decreases in serum T4 and T3

levels (Zhou et al. 2001). Similarly, acute postnatal oral exposure to pure decaBDE congener (BDE 99) 

was less potent than BDE 47, BDE 99, and 2,2',4,4',5,5’-hexaBDE (BDE 153) in inducing behavioral 

effects in adult mice (Eriksson et al. 1999, 2001a, 2002a; Viberg et al. 2001a, 2001b, 2002a, 2002b, 

2003). Chronic (103-week) exposure to a high-purity (≥97%) commercial decaBDE mixture induced 

thyroid follicular cell hyperplasia in mice, but the dose levelwas very high (2,240 mg/kg/day) and no 

thyroid histological changes occurred in similarly exposed rats (NTP 1986). 

No information is available regarding the immunosuppressive potential of PBDEs in young animals, and 

data on immune function in adult animals are mainly limited to findings in acute-duration oral studies of 

relatively high doses of pentaBDE. The plaque-forming splenic cell antibody response to injected sheep 

red blood cells was significantly reduced in mice exposed to 72 mg/kg/day pentaBDE for 14 days 

(Fowles et al. 1994), and in vitro production of IgG immunoglobulin from mitogen-stimulated 

splenocytes was reduced in mice exposed to 36 mg/kg/day pentaBDE for 14 days (Darnerud and 

Thuvander 1998). Other 14-day studies in mice found no changes in natural killer cell activity at dosages 

≤72 mg/kg/day (Fowles et al. 1994) or numbers of splenic and thymic lymphocyte subsets at dosages 

≤36 mg/kg/day (Darnerud and Thuvander 1998), although 18 mg/kg/day of the single congener 

2,2’,4,4’-tetraBDE (BDE 47) caused significantly reduced numbers of total lymphocytes and CD4+, 

CD8+, and CD45R+ subtypes in spleen (Darnerud and Thuvander 1998). Histological examinations of 

spleen, thymus, lymph node, and/or bone marrow tissues in systemic toxicity studies showed no effects of 

repeated dietary administration in intermediate-duration studies in rats exposed to ≤8,000 mg/kg/day 

decaBDE for 13 weeks (NTP 1986), in mice exposed to ≤9,500 mg/kg/day decaBDE for 13 weeks, in rats 

exposed to ≤750 mg/kg/day octaBDE for 13 weeks (IRDC 1977), or in rats exposed to ≤100 mg/kg/day 

pentaBDE for 90 days (WIL Research Laboratories 1984). Chronic ingestion of decaBDE caused lesions 

in the spleen of rats exposed to ≥1,200 mg/kg/day (splenic hematopoiesis in males) or 2,240 mg/kg/day 

(splenic fibrosis and lymphoid hyperplasia in females) for 103 weeks (NTP 1986). None of these studies 

reported changes in clinical condition that could be indicative of reduced immunocompetence. Although 

the high doses and extended exposure durations provided opportunities for the development of 

histopathology in immune system tissues and relevant clinical signs, no comprehensive immunotoxicity 

evaluations of PBDEs have been performed. 
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Information on the reproductive toxicity of PBDEs mainly consists of a one-generation study of decaBDE 

in rats that found no exposure-related functional effects (Dow Chemical Co. 1975; Norris et al. 1975a, 

1975b). No histopathological changes were observed in male or female reproductive tissues from rats 

that were exposed to PBDEs in dietary systemic toxicity studies, including decaBDE in rats at doses 

≤8,000 mg/kg/day for 13 weeks or ≤2,550 mg/kg/day for 103 weeks (NTP 1986), decaBDE in mice at 

doses ≤9,500 mg/kg/day for 13 weeks or ≤7,780 mg/kg/day for 103 weeks (NTP 1986); octaBDE in rats 

at ≤750 mg/kg/day for 13 weeks (IRDC 1977); or pentaBDE in rats at doses ≤100 mg/kg/day for 90 days 

(WIL Research Laboratories 1984). Developmental toxicity studies showed no evidence of teratogenicity 

of deca-, octa-, and pentaBDE in rats and rabbits, although fetotoxic effects, including skeletal 

ossification variations at maternally toxic doses have been observed (Argus Research Laboratories 1985b; 

Breslin et al. 1989; Dow Chemical Co. 1975, 1985; Life Science Research Israel Ltd. 1987; Norris et al. 

1975a, 1975b; WIL Research Laboratories 1986). 

Results of in vitro estrogen receptor assays suggest that there is a potential for some PBDEs to disrupt 

endocrine system functions in humans. The estrogenic and antiestrogenic activities of 17 PBDE 

congeners and 3 hydroxylated PBDEs were tested in vitro using human breast cell line assays based on 

ER-dependent luciferase reporter gene expression (Meerts et al. 2001). Eleven of 17 PBDE congeners 

showed estrogenic activity (dose-dependent luciferase induction) in the ER-CALUX assay with 

T47D.Luc cells, although the most potent PBDE congeners (2,2',4,4',6-pentaBDE > 2,4,4',6-tetraBDE 

[BDE 75] > 2,2',4,6'-tetraBDE [BDE 51] > 2,4,6-triBDE [BDE 30] > 2,3',4,4',6-pentaBDE [BDE 119]) 

had EC50 values that were 250,000–390,000 times less potent than that of the natural ligand, 17β-estradiol 

(E2). In contrast, two of the hydroxylated PBDEs, 4-(2,4,6-tribromophenoxy)phenol, 

2-bromo-4-(2,4,6-tribromophenoxy)phenol (which have bromine substitution patterns similar to the 

thyroid hormones T2 [3,5-diiodothyronine] and T3 [3,3’,5-triiodothyronine], respectively) had estrogenic 

potencies exceeding that of E2. Three of the compounds with potent estrogenic activity in the ERCALUX assay, 2,4,6-triBDE (BDE 30), 2,2',4,4',6-pentaBDE, and 4-(2,4,6-tribromophenoxy)phenol, 

were also tested for estrogenicity in ERα-specific and ERβ-specific human embryonic kidney cell lines 

(293-ERα-Luc and ERβs-Luc cells, respectively). The hydroxylated PBDE was potent in the ERαspecific cells (maximum luciferase induction similar to E2) and also showed activity in the ERβ-specific 

cells (maximum 50% induction compared to E2), whereas the ERα- and ERβ-specific cell lines were less 

responsive to BDE 30 and 2,2',4,4',6-pentaBDE. These results indicate that pure congeners of PBDEs can 

be agonists of both ERα and ERβ receptors and that metabolism to hydroxylated PBDEs may increase 

estrogenic potency. However, no estrogenic activity was demonstrated by the predominant congeners in 

biological samples, 2,2’,4,4’-tetraBDE (BDE 47) and 2,2’,4,4’,5-pentaBDE (BDE 99). The hydroxy-
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PBDE derivatives that showed minimal activity were para-substituted with a hydroxyl group, whereas the 

para position of the predominant isomers in biological samples is occupied by bromine atom, suggesting 

that the relevance of the results to commercial PBDE mixtures is unclear. Additionally, because all of the 

available estrogenicity data are in vitro, there is no evidence for estrogenicity or anti-estrogenicity of 

PBDEs or PBDE metabolites in vivo, including possible effects on reproductive development in children. 

Lower brominated PBDEs are pervasive environmental contaminants that bioaccumulate in the mother’s 

body and can be transferred to infants through the placenta and breast milk. Considering the continued 

widespread production and use of PBDEs, particularly commercial pentaBDE mixtures whose constituent 

tetra- to hexaBDE congeners are highly bioaccumulative, as well as the time lag from current-year usage 

to exposure via the food chain, it is probable that tissue concentrations among the general population will 

continue to rise. The accumulation of decaBDE in breast milk does not appear to be appreciable, in large 

part because the amount in the maternal bloodstream and available for transfer to breast milk is low due to 

poor gastrointestinal absorption and rapid elimination in the feces (Hardy 2002b). The increasing 

temporal trend for PBDEs in human tissues is illustrated by findings of an exponential increase of PBDEs 

in Swedish human breast milk from 1972 to 1997 with a doubling rate of 5 years (Norén and Meironyté 

1998, 2000). The milk concentrations of PBDEs (sum of eight congeners) on a lipid basis were 0.07 ppb 

in 1972 and 4 ppb in 1997 (Meironyté et al. 1999). Analysis of samples from 11 Finnish women showed 

that PBDE concentrations (sum of four congeners, lipid basis) were similar in breast milk and placenta, 

with ranges of 0.99–5.89 and 1.00–4.40 ppb, respectively (Strandman et al. 2000). The four highest sum 

concentrations were from women following their first childbirth. No PBPK models have been developed 

for PBDEs that could be used to quantitatively predict transfer of PBDEs via breast milk or across the 

placenta. 

Assessments of potential health risks to children were performed for commercial pentaBDE, octaBDE, 

and decaBDE products by industry as part of the EPA Voluntary Children’s Chemical Evaluation 

Program (VCCEPP) (BFRIP 2002; ENVIRON 2003a, 2003b). Based on evaluations of toxicity and 

potential exposure data, the reports concluded that none of the three commercial mixtures were expected 

to be harmful to children. These assessment documents have been independently reviewed as part of the 

VCCEP Peer Consultation process, but final reports reflecting peer review findings and EPA conclusions 

are not yet available (EPA 2004a, 2004b, 2004c). 
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5.8 BIOMARKERS OF EXPOSURE AND EFFECT 

Biomarkers are broadly defined as indicators signaling events in biologic systems or samples. They have 

been classified as markers of exposure, markers of effect, and markers of susceptibility (NAS/NRC 

1989). 

Due to a nascent understanding of the use and interpretation of biomarkers, implementation of biomarkers 

as tools of exposure in the general population is very limited. A biomarker of exposure is a xenobiotic 

substance or its metabolite(s) or the product of an interaction between a xenobiotic agent and some target 

molecule(s) or cell(s) that is measured within a compartment of an organism (NAS/NRC 1989). The 

preferred biomarkers of exposure are generally the substance itself or substance-specific metabolites in 

readily obtainable body fluid(s), or excreta. However, several factors can confound the use and 

interpretation of biomarkers of exposure. The substance being measured may be a metabolite of another 

xenobiotic substance (e.g., high urinary levels of phenol can result from exposure to several different 

aromatic compounds). Depending on the properties of the substance (e.g., biologic half-life) and 

environmental conditions (e.g., duration and route of exposure), the substance and all of its metabolites 

may have left the body by the time samples can be taken. It may be difficult to identify individuals 

exposed to hazardous substances that are commonly found in body tissues and fluids (e.g., essential 

mineral nutrients such as copper, zinc, and selenium). Biomarkers of exposure to PBBS and PBDEs are 

discussed in Section 5.8.1. 

Biomarkers of effect are defined as any measurable biochemical, physiologic, or other alteration within an 

organism that, depending on magnitude, can be recognized as an established or potential health 

impairment or disease (NAS/NRC 1989). This definition encompasses biochemical or cellular signals of 

tissue dysfunction (e.g., increased liver enzyme activity or pathologic changes in female genital epithelial 

cells), as well as physiologic signs of dysfunction such as increased blood pressure or decreased lung 

capacity. Note that these markers are not often substance specific. They also may not be directly 

adverse, but can indicate potential health impairment (e.g., DNA adducts). Biomarkers of effects caused 

by PBBS and PBDEs are discussed in Section 5.8.2. 

A biomarker of susceptibility is an indicator of an inherent or acquired limitation of an organism's ability 

to respond to the challenge of exposure to a specific xenobiotic substance. It can be an intrinsic genetic or 

other characteristic or a preexisting disease that results in an increase in absorbed dose, a decrease in the 

biologically effective dose, or a target tissue response. If biomarkers of susceptibility exist, they are 

discussed in Section 5.10 “Populations That Are Unusually Susceptible”. 
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5.8.1 Biomarkers Used to Identify or Quantify Exposure to PBBs and PBDEs 

Polybrominated Biphenyls. PBBs are environmental contaminants found mainly, but not exclusively, in 

body tissues and fluids of populations with known exposure to PBBs. Because they are lipophilic and 

have long half-lives, certain PBB congeners preferentially accumulate in lipid-rich tissues, especially 

adipose, and are present in serum and human milk. Both serum and adipose PBB levels are indicators of 

exposure, but monitoring PBBs simultaneously in samples of both types is more reliable than in serum 

only. The serum/adipose partition ratios for groups of pregnant and nonpregnant Michigan women and 

chemical workers ranged between 1:140 and 1:260; the value for Michigan male farmers was 1:325–

329 (Eyster et al. 1983). These values agree with those reported by other investigators for similar 

populations (Landrigan et al. 1979; Wolff et al. 1982). The importance of a dual determination of PBBs 

in serum and adipose can be illustrated with the following example. In a Michigan cohort, 70% of 

839 subjects were identified as having had exposure by their serum PBB levels. When adipose tissue 

results were added, an additional 24% indicated exposure (Wolff et al. 1982). The larger number of 

people with measured exposure when adipose tissue results were included reflects the higher fat content 

of adipose compared to serum and the lipophilicity of the chemicals. The partition ratio of ≈1:300 made 

the adipose limit of detection a more sensitive indicator of exposure, even though the limit of detection in 

adipose was one order of magnitude higher than in serum. Partition ratios below those reported from 

groups expected to be in equilibrium may indicate current or recent exposure (Anderson 1985). 

Using an animal physiological compartment model scaled to humans by adjusting tissue volume, blood 

flow, and clearance and rate constant parameters, it was predicted that human intake of 9.8 g of 

2,2',4,4',5,5'-hexabromobiphenyl (BB 153) from milk consumption over a 230-day period would result in 

peak tissue concentrations of 720 and 2.1 ppm in adipose and blood, respectively, at 8 months, and 

443 and 1.3 ppm, respectively (Tuey and Matthews 1980). The elimination rate after 5 years would be 

1.63 mg/day, the body burden would be 5.2 g, and the half-life would be 6.5 years. When a dose of 

0.1 mg/day for 10 months was simulated, the excretion rate in a lean individual was estimated at 

10.2 μg/day; overweight individuals had an excretion rate of 4.1 μg/day. PBB in adipose tissue from the 

lean and overweight subjects were predicted to be 2,769 and 1,103 ppb, respectively. PBB in serum 

would be 8.1 ppb in lean subjects and 3.2 ppb in overweight subjects, indicating that t1/2 increases with 

increasing fat content. These predictions point to the importance of the percentage of body fat in the 

equilibrium dynamics of PBBs and indicate that because lean individuals have a smaller fat compartment, 

all of their body tissues will have higher concentrations of PBB than those in fatter individuals of the 
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same exposure (Tuey and Matthews 1980). The assumptions on which the predictions are based do not 

reflect possible differences in fat and lean subjects due to the way that PBBs are compartmentalized 

and/or excreted as a percent of the total body burden, or in decay rates due to differential partitioning. 

As indicated above, PBBs are persistent chemicals due to their lipophilicity. Some studies have reported 

practically no change in serum PBB levels over a 12–18-month period (Wolff et al. 1979b) or over a 

3-year period (Landrigan et al. 1979). The half-life of 6.5 years predicted by Tuey and Matthews (1980) 

is shorter than half-life values determined using sera data from the Michigan PBB cohort (Blanck et al. 

2000b; Lambert et al. 1990; Rosen et al. 1995). A median half-life of 12.0 years (95% CI 4–97 years) 

was estimated based on two serum measurements from 15 women (≥20 years of age) with an initial serum 

PBB level of ≥5 ppb (Lambert et al. 1990). An analysis of 51 women (≥18.8 years of age) and 112 men 

(≥18.1 years of age) with at least two measurements 1 year apart and an initial PBB level of ≥20 ppb 

found a median half-life of 13.0 years (95% CI 6.3–infinite years) and 10.0 years (95% CI 6.7–

20.0 years), respectively (Rosen et al. 1995). Based on a median half-life of 10.8 years (95% CI 9.2–

14.7 years) for the entire group (163 persons, median PBB level 45.5 ppb), it was estimated that it will 

take more than 60 years for their PBB levels to fall below a detection limit of 1 ppb. 

Determinants of PBB serum decay were investigated in a group of 380 Michigan women (≥16 years of 

age) who had an initial PBB level of at least 2 ppb and at least two measurements taken when they were 

not pregnant (Blanck et al. 2000b). The mean initial PBB level was 20.9 ppb (standard deviation 78.7), 

and the mean time between the first and last measurement was 4.2 years (range 16.0–75.2 years). A total 

of 109 women (29%) did not have a reduction in serum PBB over time. Assuming that PBBs reached 

equilibrium in the body before substantial amounts were eliminated and before the first serum 

measurements were taken, the entire body was modeled as a single compartment for PBBs with 

exponential decay. The median PBB half-life in the entire group was 13.5 years (95% CI 10.5–

23.2 years). Subject-specific decay rate estimates were regressed on predictor variables including initial 

age, body mass index (BMI), smoking history, breast-feeding duration, and parity. The serum PBB decay 

rate was slower, resulting in a longer half-life, with higher initial PBB levels; women with initial PBB 

levels of <10 and >10 ppb had median half-lives of 12.9 and 28.7 years, respectively. The PBB decay 

rate was also slower (p=0.03) in women with an initial BMI above the median (BMI≥23). Increasing 

number of pregnancies between the first and last measurement was also associated with slower decay, but 

the effect was of borderline statistical significance (p=0.06). Breast feeding as either a continuous 

variable or as categorized by tertiles of duration (<3, 3–9, or >9 months), age, and smoking were not 

associated with serum PBB decay. 
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The average concentration of PBBs (on an adipose basis and as hexabromobiphenyl) in pooled extracts of 

several hundred individual tissue samples collected in a statistically valid manner from all nine regions of 

the continental United States was 1–2 ppb (Lewis and Sovocool 1982). Chemical workers involved in the 

PBB manufacturing process had a median adipose PBB concentration of 6,000 ppb (range 400–350, 

500 ppb); Michigan male farmers and chemical workers not involved in the PBB manufacturing process 

had a median of 1,050 ppb (range 70,000–350,000 ppb) (Eyster et al. 1983). 

Polybrominated Diphenyl Ethers. Lower brominated PBDEs are persistent environmental contaminants 

that accumulate in adipose tissue, serum, and breast milk serum of the general population, whereas 

decaBDE does not appear to be bioaccumulative. The accumulation of lower brominated congeners is 

likely due to ease of absorption, metabolism, and elimination compared to decaDBE (i.e., related to each 

molecule’s specific structure and not solely due to lipophilicity). Serum, adipose, and breast milk levels 

are indicators of exposure for lower brominated PBDEs. Lower brominated congeners in breast milk are 

useful as markers of maternal body burdens as well as lactational and in utero exposures. The 

predominant congeners identified in milk and other human tissues are 2,2',4,4'-tetraBDE (BDE 47), 

2,2'4,4',5-pentaBDE (BDE 99), and 2,2'4,4',5,5'-hexaBDE (BDE 153) (all ortho-para substituted 

congeners). Considering the continued widespread production and use of PBDEs, particularly 

commercial pentaBDE mixtures whose tetra- to hexaBDE congeners are highly bioaccumulative, as well 

as the time lag from current-year usage to exposure via the food chain, it is highly likely that tissue 

concentrations among the general population will continue to rise. The increasing temporal trend for 

PBDEs in human tissues is suggested by findings of an exponential increase of PBDEs in Swedish human 

breast milk from 1972 to 1997 with a doubling rate of 5 years (Norén and Meironyté 1998, 2000). In the 

1997 sample, the concentration of PBDE (sum of eight congeners) was 4 ppb on a lipid basis, whereas the 

1972 sample contained 0.07 ppb (Meironyté et al. 1999). PBDEs have been detected in human placenta 

at concentrations similar to those in breast milk from the same women; concentrations (sum of four 

congeners, lipid basis) ranged from 0.99 to 5.89 ppb in milk and from 1.00 to 4.40 ppb in placenta 

(Strandman et al. 2000). 

Estimates of PBDE serum concentrations among electronics-dismantling workers before and after 

exposure-free vacation (median duration 28 days, range 21–35 days) indicate that the higher brominated 

congeners have shorter half-lives than lower brominated congeners (Sjödin et al. 1999b). The median 

percentage decrease in serum concentrations, based on 5–11 measurements per congener, were 14 (range 

3.5–39), 14 (2.1–38), 14 (6.7–42), 30 (7.9–52), and 66 (47–100) for 2,2',4,4'-tetraBDE (BDE 47), 
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2,2',4,4',5,5'-hexaBDE (BDE 153), 2,2',4,4',5,6'-hexaBDE (BDE 154), 2,2',3,4,4',5',6-heptaBDE 

(BDE 183), and 2,2',3,3',4,4',5,5',6,6'-decaBDE (BDE 209), respectively. Although actual half-lives were 

not calculated, the data suggest that the half-lives of the lower brominated congeners were <1 year. 

Congener patterns in humans may provide information on the nature or pathway of PBDE exposures 

(Hooper and McDonald 2000). Low tetra:deca congener ratios are suggestive of direct, recent, or 

occupational exposures to the parent PBDE mixture, whereas higher ratios may indicate an environmental 

pathway where exposures result from PBDEs that have leached from the parent mixtures and have been 

degraded in the environment. 

5.8.2 Biomarkers Used to Characterize Effects Caused by PBBs and PBDEs 

Polybrominated Biphenyls. Biomarkers of effects for PBBs are likely to be common to the general class 

of halogenated aromatic hydrocarbons, rather than specific for PBBs, because PCBs, PBDEs, and other 

structurally similar chemicals cause generally similar effects. A potential biomarker for PBBs is related 

to their effect on the thyroid gland. As discussed in Sections 5.2.2.2, Endocrine Effects, the thyroid gland 

is a sensitive and unequivocal target of PBBs in animals, and evidence in humans is suggestive of a 

similar relationship. Effects in workers exposed to unspecified PBBs and/or decabromobiphenyl included 

increased serum thyrotropin, low or borderline low serum T4, and increased thyroid antimicrosomal 

antibody titers (Bahn et al. 1980). A spectrum of thyroid effects has been observed in exposed rats, 

ranging from decreases in serum levels of serum T4 and T3 to histological and ultrastructural changes 

(Akoso et al. 1982b; Allen-Rowlands et al. 1981; Byrne et al. 1987; Gupta and Moore 1979; Kasza et al. 

1978a; Meserve et al. 1992; Norris et al. 1975a, 1975b; NTP 1983; Sepkovic and Byrne 1984; Sleight et 

al. 1978; Werner and Sleight 1981). Therefore, serum levels of T4 and/or other thyroid hormones are 

potential biomarkers of effects for PBBs. Additional studies could better characterize thyroid effects of 

PBBs in humans and develop specific correlations between levels and duration of exposure and 

alterations in serum T4 and T3 levels, including information on the specific amount of change in the 

biomarkers associated with a demonstrably adverse effect. These potential biomarkers are not specific to 

PBBs because PBDEs and other antithyroid agents can have similar effects. 

Caffeine has been used as a potential tool to characterize exposure and/or effect of PBBs (Lambert et al. 

1990). In this test, caffeine is used as a metabolic probe of cytochrome P-450 isozymes activity from the 

CYP1A family, which is significantly induced by PBBs in animals (Safe 1984). The caffeine breath test 

(CBT) is primarily useful for detecting induction of CYP1A2 activity in human liver. Because the 
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induction of CYP1A enzymes is AhR mediated, the test has been used as a marker for exposure to PCBs, 

CDDs, and CDFs (Lambert et al. 1992). A volunteer population of 50 Michigan subjects with previously 

high serum PBB levels and 50 with undetectable or low serum levels was compared to a control 

population not exposed to PBBs (Lambert et al. 1992). Two tests were conducted, the CYP1A2-

dependent caffeine 3-N-demethylase activity was monitored by the CBT, and 7-N-demethylase activity 

was monitored by the caffeine urinary metabolite ratio (CMR). PBB-exposed subjects had higher CBT 

values (p<0.02) than urban nonsmokers, but the values were comparable to those of urban smokers. The 

correlation between serum PBB levels and the CBT value was poor (r2

=0.2). The CMR value in PBBexposed subjects was also higher than that of urban nonsmokers (p<0.05); there was no correlation 

between serum PBB levels and CMR values. Generally, smokers have higher CBT values than 

nonsmokers due to the presence of polycyclic aromatic hydrocarbons (PAH) in tobacco smoke, which 

induce CYP1A (Kotake et al. 1982). 

Many reports have been published regarding possible associations between PBB exposure and adverse 

health effects in populations from the state of Michigan. An early study compared the health status of 

people on quarantined farms with people in nonquarantined farms in the same area (Humphrey and 

Hayner 1975). The results showed no pattern of differences between the groups. Moreover, no 

abnormalities of heart, liver, spleen, or nervous system that could be related to PBB exposure were found 

in physical examinations. A follow-up study examined the prevalence of selected symptoms in groups of 

varying potential exposure 4 years after exposure (Landrigan et al. 1979). In general, symptoms were 

more prevalent in two self-selected groups and were least prevalent in the group composed of chemical 

workers. No positive associations were found between serum PBB concentrations and symptom 

frequencies; yet a third group of studies reported an increased incidence of symptoms in Michigan 

farmers relative to a group of control Wisconsin farmers (Anderson et al. 1978a, 1978b, 1978c, 1979). As 

observed in other epidemiology studies, self-selected groups, which had lower PBB concentrations in 

serum, reported a high incidence of symptoms, compared to randomly selected groups. No specific 

biomarker of effect could be identified in the Michigan contamination episode. Furthermore, the 

prevalence of the reported symptoms had no consistent relationship to the extent or types of exposure, and 

most objective clinical measurements have failed to show a significant relationship to PBB exposure 

(Fries 1985a). Additional information regarding biomarkers for effects can be found in OTA (1990) and 

CDC/ATSDR (1990). For a more detailed discussion of the health effects caused by PBBs see 

Section 5.2. 
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Polybrominated Diphenyl Ethers. Biomarkers of effects for PBDEs are likely to be common to the 

general class of halogenated aromatic hydrocarbons, rather than specific for PBDEs, because PBBs, 

PCBs, and other structurally similar chemicals cause generally similar effects. The thyroid is a critical 

target for lower brominated PBDEs in animals. As discussed in Sections 5.2.2.2, Endocrine Effects, 

exposure to commerical penta- and octaBDE products caused thyroid changes in rats and mice, 

particularly reduced serum T4 levels (Darnerud and Sinjari 1996; Fowles et al. 1994; Hallgren and 

Darnerud 1998; WIL Research Laboratories 1984; Zhou et al. 2001, 2002). The potential for induction of 

thyroid effects by decaBDE appears to be low in comparison to penta- and octaBDE mixtures due to 

toxicokinetic and toxicity differences. In particular, lower brominated BDE congeners preferentially 

accumulate in the body compared to decaBDE, apparently due to their greater partitioning and retention 

in lipid-rich tissues and lower rates of metabolism and elimination (Hardy 2002b; Morck et al. 2001, 

2003), and decaBDE only affected the thyroid in animals following lifetime exposure to dose levels 

several orders of magnitude higher than effect levels of penta- and octaBDE in acute-duration studies 

(NTP 1986; Zhou et al. 2001). 

Serum T4 may not be a reliable biomarker in humans because the extent that PBDEs affect circulating 

levels of T4 (or T3) possibly varies with species. As discussed in Section 5.5.3, Animal-to-Human 

Extrapolations, humans are generally regarded as less sensitive than rats to thyroid effects of PBDEs due 

to differences in the binding of T4 to serum transport proteins. In particular, in humans, the binding of 

T4 to thyroxine binding globule (TBG) protects the hormones from metabolism and excretion, resulting 

in relatively long serum half-lives. Rats lack a high-affinity binding protein (i.e., TBG) and consequently 

have lower serum levels of T4 due to increased availability of the hormone for metabolism and 

elimination. The clearance of T4 in rats is enhanced by the induction hepatic microsomal enzymes, 

particularly UDP-glucuronyl transferase, which catalyzes the conjugation of free T4 and enhances its 

excretion in the bile. Effects of PBDEs on thyroid status via induction of hepatic enzymes are therefore 

unlikely to occur in humans. However, because PBDE-induced decreases in T4 appear to involve 

multiple mechanisms in addition to increased thyroid hormone metabolism (Zhou et al. 2001), serum T4

level is a potential biomarker of effects for lower brominated PBDEs in humans. Additional studies 

would be useful to characterize the applicability of the biomarker to humans and develop possible 

correlations between levels and duration of exposure and changes in serum T4. This biomarker is not 

specific to PBDEs because other antithyroid agents can have similar effects. Additionally, screening of 

newborn children for hypothyroidism is a widely accepted and legislatively mandated practice (LaFranchi 

1999; Landenson et al. 2000). 
PBBs and PBDEs 310 

5. HEALTH EFFECTS 

5.9 INTERACTIONS WITH OTHER CHEMICALS 

Polybrominated Biphenyls. PBBs are potent inducers of liver and kidney P-450 enzymes (MFO) (Haake 

et al. 1985; Halvorson et al. 1985; Shepherd et al. 1984), and as such, they could potentially enhance or 

decrease the toxicity of any substance that is metabolized by the P-450 system. PBBs are thought to 

potentiate the hepatotoxicity and nephrotoxicity of halogenated hydrocarbons and other substances by 

inducing P-450s that biotransform them to more toxic metabolites (Ahmadizadeh et al. 1984; Kluwe and 

Hook 1978; Kluwe et al. 1978, 1979, 1982; Kuo and Hook 1982; Roes et al. 1977). In these studies, rats 

and/or mice were given diets containing FireMaster BP-6 that provided doses of 0.13–13 mg/kg/day for 

periods of 10–28 days prior to intraperitoneal challenge with the halogenated hydrocarbons. 

Nephrotoxicity was assessed by measuring kidney weights and the levels of blood urea nitrogen, and by 

the accumulation of p-aminohippurate and/or tetraethylammonium (TEA) in renal cortical slices. 

Hepatotoxicity was assessed by relative liver weights and by levels of SGPT and/or SGOT. In most 

cases, exposure to PBBs alone did not affect the parameters of nephrotoxicity in animals. However, 

exposure to PBBs alone usually caused increased relative liver weights and elevated levels of SGOT and 

SGPT. Pre-exposure to PBBs increased the hepatotoxicity and nephrotoxicity of chloroform 

(Ahmadizadeh et al. 1984; Kluwe and Hook 1978; Kluwe et al. 1978) and carbon tetrachloride (Kluwe et 

al. 1979, 1982) and the nephrotoxicity of trichloroethene and 1,1,2-trichloroethane (Kluwe et al. 1978, 

1979). PBB pretreatment in dietary studies also potentiated the nephrotoxicity of the antibiotic, 

cephaloridine, in rats (Kuo and Hook 1982). 

Pretreatment with PBBs also potentiated the lethality of chloroform, carbon tetrachloride, and 1,1,2-trichloroethane by decreasing the LD50 values (Kluwe et al. 1978, 1979) and the lethality of bromobenzene 

by decreasing the time to death (Roes et al. 1977) in mice after challenge with the halogenated hydrocarbon. In contrast, pretreatment of mice with PBBs in dietary studies increased the LD50 value of 1,2-dibromo-3-chloropropane (DBCP) but had no effect on the LD50 value of 1,2-dibromoethane (EDB) (Kluwe 

et al. 1981). Also, PBBs were found to reverse the depletion of nonprotein sulfhydryls (e.g., glutathione) 

caused by DBCP and EDB in the livers and kidneys of mice, suggesting that PBB exposure protected the 

mice from the lethality of DBCP by making glutathione more available for conjugation with the toxic 

metabolites. 

No potentiation of toxicity was found when rats were co-exposed to diets containing PBB and mirex, 

photomirex, or kepone, compared with toxicity elicited by each of these substances alone (Chu et al. 

1980). 
PBBs and PBDEs 311 

5. HEALTH EFFECTS 

Polybrominated Diphenyl Ethers. No specific information was located regarding interactions between 

PBDEs and other chemicals. PBDEs are a group of 209 congeners that display biological activity 

involving different potential mechanisms, and share some toxicological properties with structurally 

similar polyhalogenated aromatic compounds, particularly PBBs, PCBs, polychlorinated- and polybrominated dibenzo-p-dioxins (PCDDs and PCBDs), and chlorinated and brominated dibenzofurans 

(PCDFs and PBDFs). However, although these chemicals are structurally similar in two dimensions, 

PBDEs differ from the other classes on a three-dimensional basis. In particular, as discussed 

Section 5.5.2 Mechanisms of Toxicity, the influences of the ether bridge in the PBDE molecule preclude 

clearly classifying the congeners as either dioxin-like (coplanar) or non-dioxin-like (non-coplanar). This 

has implications for both dioxin-type toxicities (mediated by the AhR pathway) and non-dioxin-type 

effects. For example, structure-activity studies of PBDEs have shown that binding affinities and 

induction of AhR-mediated responses are very weak or negligible (Chen et al. 2000; Meerts et al. 1998), 

suggesting that there is a low potential for effects and interactions involving Ah-receptor-mediated 

mechanisms (e.g., induction of hepatic CYP1A oxygenases and Phase II enzymes, body wasting, thymic 

atrophy, and porphyria). Because of the diversity of PBDE activities involving Ah-receptor-independent 

mechanisms (e.g., induction of CYP2B and CYP3A oxygenases, induction of changes in brain dopamine 

levels, and disruption of calcium homeostasis), or possibly both Ah-receptor-dependent and -independent 

mechanisms (e.g., liver hypertrophy, disruption of steroid hormone homeostasis, thyroid hormone 

homeostasis, disruption of immune functions), there is a large potential for PBDEs to alter the toxicity of 

other chemicals, or other chemicals to alter the toxicity of PBDEs. For example, lower brominated PBDE 

mixtures and congeners are inducers of hepatic microsomal enzymes (Carlson 1980a, 1980b; Fowles et al. 

1994; Hallgren et al. 2001; Zhou et al. 2001, 2002) and therefore could potentially enhance or decrease 

the toxicity of any substance that is metabolized by the P-450 system. 

5.10 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE 

A susceptible population will exhibit a different or enhanced response to PBBs and PBDEs than will most 

persons exposed to the same level of PBBs and PBDEs in the environment. Reasons may include genetic 

makeup, age, health and nutritional status, and exposure to other toxic substances (e.g., cigarette smoke). 

These parameters result in reduced detoxification or excretion of PBBs and PBDEs, or compromised 

function of organs affected by PBBs and PBDEs. Populations who are at greater risk due to their 

unusually high exposure to PBBs and PBDEs are discussed in Section 8.7, Populations with Potentially 

High Exposures. 
PBBs and PBDEs 312 

5. HEALTH EFFECTS 

Information was located on a small part of the U.S. population that might be unusually susceptible to 

PBBs and PBDEs. As indicated in Section 5.4.4.2, breast milk constitutes the most important route of 

excretion of PBBs in lactating females, and this is likely to be the case for lower brominated PBDEs as 

well. DecaBDE has not been reported in breast milk and is unlikely to accumulate in the milk due to poor 

absorption and rapid elimination in the feces (Hardy 2002b). Therefore, women with high body burdens 

of PBBs and/or lower brominated PBDEs who breast-feed may be placing their infants at a higher risk of 

potential health effects. However, in most cases, the benefits of breast-feeding are expected to outweigh 

any risk to infants from exposure to these chemicals in the maternal milk. 

Experiments in animals and model simulations in humans have shown that reduction in body fat markedly 

decreases the elimination half-life of PBBs (Domino et al. 1982; Tuey and Matthews 1980). For 

example, when a dose of 0.1 mg/day for 10 months was simulated in humans, the excretion rate in a lean 

individual was estimated at 10.2 μg/day; overweight individuals had an excretion rate of 4.1 μg/day. The 

cumulative excretion was 51% of the dose in lean subjects compared to 20.7% in overweight subjects. 

These data indicate that overweight individuals may be at higher risk because they store PBBs for a 

longer time than lean subjects. On the other hand, because lean individuals have a smaller fat 

compartment, their body tissues will contain higher concentrations of PBB than those in subjects with 

more fat who received the same exposure (Tuey and Matthews 1980); thus, leaner individuals may be 

more vulnerable to short-term effects than fatter individuals. Because of this phenomenon, a sudden 

reduction in body fat, such as that which could occur during dieting, may cause a redistribution of PBBs 

to potential target organs, which would also increase the potential for adverse health effects to such 

individuals. 

Pregnant women and developing infants and fetuses should be viewed as possibly sensitive populations 

for exposure to PBBs and lower brominated PBDEs as they are for other thyroid hormone disrupting 

chemicals (Glinoer 1990; McDonald 2002; Morreale de Escobar et al. 2002). The condition of pregnancy 

normally puts a significant strain on the maternal thyroid system, which can be exacerbated by iodine 

deficiency; according to data from 1988 to 1994, iodine deficiency is prevalent in approximately 12% of 

the general population and 15% of women of child-bearing age in the United States (Hollowell et al. 

1998). Thyroid hormones are essential for normal development of the nervous system, lung, skeletal 

muscle, and possibly other organ systems, and the fetus is dependent on maternal thyroid hormones at 

least until the fetal thyroid begins to produce T4 and T3, which occurs in humans at approximately 16–

20 weeks of gestation (Zoeller and Crofton 2000). The potential of lower brominated PBDEs to disrupt 

maternal, fetal, and newborn thyroid hormone levels is demonstrated by the Zhou et al. (2001, 2002) 
PBBs and PBDEs 313 

5. HEALTH EFFECTS 

studies of pentaBDE (technical mixture) in rats, as discussed in Section 5.2.2.2 (Endocrine Effects 

subsection). The potential for induction of thyroid and neurodevelopmental effects by decaBDE is low in 

comparison to commercial penta- and octaBDE mixtures due to toxicokinetic and toxicity differences, as 

discussed in .Section 5.7 (Children’s Susceptibility). 

People with exposure to anti-thyroid drugs (e.g., lithium), thyroid disease, or otherwise compromised 

thyroid function might have a more pronounced response to PBBs and PBDEs because of their underlying 

limitations in thyroid hormone production. Similarly, people with compromised function of other organs, 

such as those with liver or kidney diseases (e.g., liver cirrhosis or hepatitis B), could be considered more 

susceptible to health effects of PBBs and PBDEs. However, the potential for induction of thyroid effects 

by PBDEs in humans appears to be low, based on mechanistic data as discussed Section 5.5.3 (Animal-toHuman Extrapolations). 

5.11 METHODS FOR REDUCING TOXIC EFFECTS 

This section will describe clinical practice and research concerning methods for reducing toxic effects of 

exposure to PBBs and PBDEs. However, because some of the treatments discussed may be experimental 

and unproven, this section should not be used as a guide for treatment of exposures to PBBs and PBDEs. 

When specific exposures have occurred, poison control centers and medical toxicologists should be 

consulted for medical advice. Specific treatment methods have not been developed for exposures to PBBs 

or PBDEs. 

The treatment methods discussed below are general methods that would apply to any persistent, lipophilic 

chemical, and have not been tested for efficacy, indicating that they might not be effective in reducing the 

toxic effects of PBBs and PBDEs. There is no indication of hazards associated with the treatments. The 

methods are particularly appropriate for trying under conditions of acute exposure, but PBBs and PBDEs 

are not acutely toxic chemicals. Scenarios where life-threatening acute exposure would occur are 

unlikely, although accidental or intentional ingestion of the commercial products is a conceivable 

concern. The relevance of the methods to common background environmental exposures to these 

chemicals is unclear, and it is questionable whether current exposure and tissue levels in the general 

population are a health concern. 
PBBs and PBDEs 314 

5. HEALTH EFFECTS 

5.11.1 Reducing Peak Absorption Following Exposure 

Ingested PBBs and PBDEs are absorbed by the gastrointestinal tract of humans and animals (see 

Section 5.4). Although there are no specific recommendations for clinical treatment of acute intoxication 

from ingested PBBs and PBDEs, recommendations based on experiences with PCBs are relevant. 

Treatments for acute poisonings from PCBs and related substances include the induction of emesis or 

gastric lavage and stomach pumping to decrease gastrointestinal absorption of the chemicals (Lemesh 

1992). These procedures would not be beneficial if performed too long after exposure occurred. 

Administration of activated charcoal as a slurry, either aqueous or mixed with a saline cathartic or 

sorbitol, is frequently recommended to decrease the gastrointestinal absorption of PCBs, but the value of 

this treatment for reducing absorption of PCBs, PBBs, and PBDEs is unknown (HSDB 1992). Repetitive 

administration of activated charcoal might be useful in preventing reabsorption of metabolites. Rice bran 

fiber decreased absorption of PCBs in the gastrointestinal tract and had a stimulatory effect on fecal 

excretion of PCBs in rats (Takenaka et al. 1991), but it is unclear if rice bran would be of benefit in 

poisoned humans. 

The detection of PBBs and PBDEs in the serum and fat of people who were occupationally exposed to 

these chemicals indicates that PBBs and PBDEs can be absorbed by the lungs,skin, and/or orally by handto-mouth contact. Although no specific methods to reduce absorption of dermally applied or inhaled 

PBBs or PBDEs were located, multiple washings of contaminated skin with soap and water immediately 

following exposure have been suggested to reduce the dermal absorption of PCBs (HSDB 1992). Studies 

with monkeys showed that soap and water was as effective as or better than such solvents as ethanol, 

mineral oil, or trichlorobenzene in removing PCBs from skin (Wester et al. 1990). Personal protective 

equipment (e.g., long sleeves, gloves, safety glasses, respiratory protection) and industrial hygiene 

programs generally help to limit occupational exposures. 

5.11.2 Reducing Body Burden 

PBBs and lower brominated PBDEs tend to accumulate in lipid-rich tissues and are slowly metabolized 

and eliminated from the body (see Section 5.4). DecaBDE differs from the lower brominated PBDEs in 

that it is unlikely to preferentially accumulate in body tissues due to poor oral absorption and metabolism 

and rapid elimination in the feces (Hardy 2002b). Several methods to enhance the elimination of PBBs 

from the body have been examined in animals and are applicable to PBDEs, although the relevance of the 

methods is questionable because it is unclear whether current tissue levels are a health concern for the 
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5. HEALTH EFFECTS 

general population. Methods for increasing the elimination of these chemicals include the restriction of 

caloric intake (to reduce total body fat), and the administration of various agents that interact with bile 

acids including activated charcoal, mineral oil and bile-binding resins such as cholestyramine 

(Kimbrough et al. 1980; McConnell et al. 1980; Polin and Leavitt 1984; Polin et al. 1985, 1991; Rozman 

et al. 1982). It should be mentioned, however, that based on the pharmacokinetic considerations 

discussed in Section 5.8.1, a rapid breakdown of fat, as might occur in dieting, might lead to a transient 

increase in PBB and PBDE levels in serum and other body tissues, possibly posing a significant reexposure problem. Although some of the studies observed no enhanced elimination (Kimbrough et al. 

1980; McConnell et al. 1980), others identified treatments that were effective in enhancing the biliary and 

intestinal elimination of PBB residues (Polin et al. 1991; Rozman et al. 1982). Polin et al. (1991) found 

that dietary intervention to reduce PBBs was dose dependent; treatment with 10% mineral oil and a 45% 

reduction in food intake resulted in a 69 and 23% reduction in body burden in rats fed PBBs at dietary 

concentrations of 0.1 and 100 ppm, respectively (Polin et al. 1991). A combination of mineral oil, 

colestipol, and dietary restriction was successful in reducing the PBB body burdens in chickens (Polin 

and Leavitt 1984; Polin et al. 1985), while each treatment alone had no effect in reducing PBB body 

burden. A 3-week treatment regimen that included dietary supplements of polyunsaturated oil, vitamins, 

and minerals, and heat stress has been applied in a pilot study to seven human subjects that were known to 

have been exposed to PBBs; following treatment, statistically significant reductions were measured in 

PBB concentrations in fat (Schnare et al. 1984). Although the lack of a separate control group 

complicates interpretation of the results of this study (each subject served as his/her own control), this 

treatment was developed for the purpose of reducing body burdens of fat-soluble psychoactive drugs 

(Schnare et al. 1984). A liquid diet was used for 16 individuals who developed symptoms following 

exposure to PCBs and polychlorinated dibenzofurans (Imamura and Tung 1984). Symptoms were 

reduced several months after the fasting period. This study is limited in that a control group was not used, 

and body burdens were not measured. Based on information for PCBs, mobilization of PBBs or PBDEs 

from adipose tissue is not recommended in individuals with hepatic or renal disease (Lemesh 1992). 

5.11.3 Interfering with the Mechanism of Action for Toxic Effects 

There are no known methods for interfering with the mechanism of action of PBBs or PBDEs. Although 

the mechanism of action of PBBs and PBDEs is not completely understood, experimental evidence 

accumulated in recent years indicates that some PBB congeners exert toxic actions by a process involving 

several steps (Safe 1984). This process begins with the binding of particular congeners to the AhR and 

leads ultimately to enhancement of the CYP1A1 gene expression (see Section 5.5). It appears, therefore, 
PBBs and PBDEs 316 

5. HEALTH EFFECTS 

that interfering with the initial step, binding to the receptor, or with any of the subsequent steps, would 

possibly prevent the expression of the toxic effects. Several compounds have been identified that 

partially antagonize one or more AhR-mediated responses (Bannister et al. 1989); their use, however, has 

been limited to experimental studies in animals. These compounds were successful antagonists when 

given before or at the same time as an AhR activator (2,3,7,8-tetrachlorodibenzo-p-dioxin) (Bannister et 

al. 1989). Therefore, the potential for interfering with the mechanism of AhR-mediated effects of PBBs 

after exposure has occurred is largely untested. For PBDEs, the ether bridge precludes congeners from 

readily adopting the coplanar (dioxin-like) conformation characteristic of AhR ligands (see Section 5.5.2 

Mechanisms of Toxicity). Because PBDE congeners are not clearly classifiable as either dioxin-like 

(coplanar) or non-dioxin-like (non-coplanar), AhR antagonists might not effectively antagonize effects of 

PBDEs. 

PBBs and PBDEs may also cause toxicity by other mechanisms of action. For example, some PBB 

congeners can be metabolized to reactive arene oxides (Kohli and Safe 1976; Kohli et al. 1978) that may 

alkylate critical cellular macromolecules and result in injury. PBDE-induced decreases in thyroid T4

hormone, which can affect neurobehavioral development, are likely to involve multiple mechanisms (see 

Section 5.5.3). These include induction of hepatic microsomal enzymes, particularly UDPGT, which can 

increase the rate of T4 conjugation and excretion, and metabolic formation of hydroxy-metabolites of 

PBDEs. PBDEs and their hydroxy metabolites can bind with high affinity to thyroid transport proteins 

because they are structurally similar to T4 hormone (i.e., are also hydroxy-halogenated diphenyl ethers) 

(see Section 5.5.2). Effects of PBDEs on thyroid status via induction of hepatic enzymes, however, are 

unlikely to occur in humans, and the impact of hydroxy-metabolites on serum T4 needs further 

clarification. Effects of PBDEs on the function and development of the nervous system could also 

involve disruption of calcium homeostatic mechanisms and intracellular signalling events (Kodavanti 

2003; Kodavanti and Derr-Yellin 2001, 2002; Smolnikar et al. 2001; Wiegand et al. 2001), effects on 

dopamine uptake into synaptic vesicles and synaptosomes (Mariussen and Fonnum 2002, 2003; 

Mariussen et al. 2003a) and cholinergic functions (Viberg et al. 2001b, 2002a, 2002b), and/or free radicalinduced neuronal death (Reistad et al. 2002). Clinical interventions designed to interfere with the 

aforementioned mechanisms have yet to be developed. 

5.12 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the 

Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 
PBBs and PBDEs 317 

5. HEALTH EFFECTS 

adequate information on the health effects of PBBs and PBDEs is available. Where adequate information 

is not available, ATSDR, in conjunction with the National Toxicology Program (NTP), is required to 

assure the initiation of a program of research designed to determine the health effects (and techniques for 

developing methods to determine such health effects) of PBBs and PBDEs. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will be 

evaluated and prioritized, and a substance-specific research agenda will be proposed. 

5.12.1 Existing Information on Health Effects of PBBs and PBDEs 

The existing data on health effects of inhalation, oral, and dermal exposure of humans and animals to 

PBBs and PBDEs are summarized in Figures 5-6 and 5-7. The purpose of this figure is to illustrate the 

existing information concerning the health effects of PBBs and PBDEs. Each dot in the figure indicates 

that one or more studies provide information associated with that particular effect. The dot does not 

necessarily imply anything about the quality of the study or studies, nor should missing information in 

this figure be interpreted as a “data need”. A data need, as defined in ATSDR’s Decision Guide for 

Identifying Substance-Specific Data Needs Related to Toxicological Profiles (Agency for Toxic 

Substances and Disease Registry 1989), is substance-specific information necessary to conduct 

comprehensive public health assessments. Generally, ATSDR defines a data gap more broadly as any 

substance-specific information missing from the scientific literature. 

As indicated in Figure 5-6, information is available regarding systemic, immunological, neurological, 

developmental, reproductive, and carcinogenic effects of PBBs in humans. The information on effects in 

humans is derived from the Michigan contamination episode that involved chronic-duration oral exposure 

to contaminated food and from occupational exposure data in which it was assumed that exposure was 

predominantly through skin contact, although inhalation exposure cannot be ruled out. 

Information on health effects in animals is extensive and available for all effect categories, but is nearly 

completely limited to oral exposure studies, which appears to reflect experimental practicality and 

concern for what is thought to be the most prevalent and likely route of environmental exposure. 
PBBs and PBDEs 318 

5. HEALTH EFFECTS 

Figure 5-6. Existing Information on Health Effects of PBBs 

Existing Studies

Inhalation

Oral

Dermal

Death

Acute

Intermediate

Chronic

Immunologic/Lymphoretic

Neurologic

Reproductive

Developmental

Genotoxic

Cancer

Systemic

Animal

Inhalation

Oral

Dermal

Death

Acute

Intermediate

Chronic

Immunologic/Lymphoretic

Neurologic

Reproductive

Developmental

Genotoxic

Cancer

Systemic

Human
PBBs and PBDEs 319 

5. HEALTH EFFECTS 

Figure 5-7. Existing Information on Health Effects of PBDEs 

Existing Studies

Inhalation

Oral

Dermal

Death

Acute

Intermediate

Chronic

Immunologic/Lymphoretic

Neurologic

Reproductive

Developmental

Genotoxic

Cancer

Systemic

Animal

Inhalation

Oral

Dermal

Death

Acute

Intermediate

Chronic

Immunologic/Lymphoretic

Neurologic

Reproductive

Developmental

Genotoxic

Cancer

Systemic

Human
PBBs and PBDEs 320 

5. HEALTH EFFECTS 

A limited amount of information is available on the systemic and carcinogenic effects of PBDEs in 

humans (Figure 5-7). Information on health effects of PBDEs in animals is available for all effect 

categories but, like PBBs, is mainly limited to oral exposure studies in animals. In general, the health 

effects of PBDEs are less adequately studied than for PBBs (and PCBs), and a prominent limitation of the 

PBDE database is a lack of adequate human studies. 

5.12.2 Identification of Data Needs 

Acute-Duration Exposure. 

Polybrominated Biphenyls. The hepatotoxicity of PBBs in rats and mice is reasonably well characterized 

for acute-duration oral exposure (Bernert et al. 1983; Corbett et al. 1975; Gupta and Moore 1979; Gupta 

et al. 1981; Kimbrough et al. 1978b, 1980, 1981; Lee et al. 1975a, 1975b; Norris et al. 1975a; Raber and 

Carter 1986; Waritz et al. 1977). Effects on body weight in rats and mice and on the thyroid in rats are 

also well documented (Allen-Rowlands et al. 1981; Corbett et al. 1978; Fraker 1980; Gupta and Moore 

1979; Kimbrough et al. 1981), and thyroid effects occurred at doses as low as those causing liver effects. 

Insufficient acute data exist to definitely establish if the thyroid effects are more critical than effects in the 

liver, but extensive data on thyroid effects from longer term studies and the functional nature of the 

changes suggest that this is the case and justify using a thyroid effect as the basis for an acute oral MRL. 

Acute oral studies in other species would clearly establish the most sensitive target and species for acute 

exposure. Tests with monkeys, guinea pigs, and mink would be informative because intermediate- and 

chronic-duration studies indicate that these species are more sensitive than the rat and that 

endocrinological effects are particularly sensitive end points. 

Information on toxic effects of acute-duration exposure to PBBs by routes other than oral are limited to 

data on hepatic, renal, dermal, and ocular effects of inhalation and dermal exposure in rats or rabbits 

(Millischer et al. 1980; Needham et al. 1982; Norris et al. 1975a; Waritz et al. 1977), but these data may 

not be reliable due to study limitations and possible delayed lethality. Limitations in the animal database 

include inadequate reporting (e.g., numbers of animals not reported), limited number of exposure levels, 

and lack of studies of PBB mixtures likely to be most toxic (i.e., Firemaster PBBs). Quantitative data for 

inhalation and dermal absorption of PBBs are lacking. Studies of inhalation and dermal absorption 

following exposure to soil containing PBBs (i.e., bioavailability studies) would be useful for assessing 

risk at a hazardous waste site. Further studies identifying target organs and examining the dose-response 
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5. HEALTH EFFECTS 

relationship following acute inhalation and dermal exposure to PBBs would also be informative, although 

exposure via soil and acute toxicosis is not likely to ever be a concern. 

Polybrominated Diphenyl Ethers. Acute-duration studies have documented effects of PBDEs mainly on 

the liver and thyroid of orally exposed rats and mice (Argus Research Laboratories 1985a; Carlson 1980a, 

1980b; Darnerud and Sinjari 1996; Fowles et al. 1994; Hallgren et al. 2001; IRCD 1974, 1975a, 1975b; 

Life Science Research Israel Ltd. 1987; NTP 1986; WIL Research Laboratories 1986; Zhou et al. 2001, 

2002). The available data on lower brominated PBDEs indicate that the thyroid is a particularly sensitive 

target of acute oral exposure and justify using thyroid effects as the basis for an acute oral MRL, but acute 

effects of lower brominated PBDEs on the liver are not as well characterized as thyroid effects. Other 

studies indicate that immunosuppression and neurobehavior are important and potentially critical health 

end points for acute exposure to lower brominated PBDEs that need to be further investigated (see 

discussions of data needs for Immunotoxicity and Neurotoxicity). Studies in other species would help to 

clearly establish the most sensitive target and species for acute exposure, as well as which animal toxicity 

data are the most relevant to humans and useful for assessing acute health risks of lower brominated 

PBDEs.. 

Data on decaBDE are insufficient for derivation of an acute-duration oral MRL. Information is available 

on effects of acute oral exposure to decaBDE on body and liver weights, microsomal enzyme induction in 

the liver, and serum thyroid levels in weanling rats (Carlson 1980b; NTP 1986; Zhou et al. 2001), but the 

database is limited by lack of LOAELs and/or sufficiently sensitive end points for estimation of an MRL. 

Acute-duration inhalation exposure toxicity studies of decaBDE were not located. The inhalation 

database for acute-duration exposure to lower brominated PBDEs is essentially limited to two 14-day 

unpublished industry-sponsored studies of octaBDE in rats (Great Lakes Chemical Corporation 1978, 

2000). Liver and nasal effect levels were identified in these studies, but MRL estimation is precluded by 

inconsistencies between the studies and a lack of information on thyroid hormone levels. Additional 

dose-response studies are needed to provide an adequate basis for derivation of an acute inhalation MRL 

for lower brominated PBDEs. 

Intermediate-Duration Exposure. 

Polybrominated Biphenyls. The preponderance of toxicity data for PBBs are available from animals 

exposed to FireMaster FF-1 or FireMaster BP-6 in the diet or by gavage in intermediate-duration studies 

(Akoso et al. 1982a, 1982b; Allen et al. 1978; Allen-Rowlands et al. 1981; Aulerich and Ringer 1979; 
PBBs and PBDEs 322 

5. HEALTH EFFECTS 

Byrne et al. 1987, 1988; Castracane et al. 1982; Darjono et al. 1983; Gupta and Moore 1979; Gupta et al. 

1981; Kasza et al. 1978a, 1978b; Ku et al. 1978; Lambrecht et al. 1978; Lee et al. 1975a, 1975b; Loose et 

al. 1981; McCormack et al. 1978; Norris et al. 1975a; NTP 1983; Ringer et al. 1981; Sepkovic and Byrne 

1984; Sleight and Sanger 1976; Sleight et al. 1978; Waritz et al. 1977; Werner and Sleight 1981). Studies 

have been performed with various species (rats have been tested most extensively), and there is evidence 

indicating that monkeys, guinea pigs, and mink may be the most sensitive. The liver, skin, stomach, and 

thyroid are unequivocal targets, but existing studies do not identify NOAELs for toxic effects in these 

organs in rats and/or more sensitive species. Hematologic changes indicative of anemia are consistently 

reported effects in various species, but the relative importance of these effects is not known. Evidence 

suggests that the LOAELs for thyroid effects in rats and hepatic effects in guinea pigs are similar (Akoso 

et al. 1982b; Sleight and Sanger 1976), but reproductive and developmental effects occurred in monkeys 

at a lower dosage. The serious nature of the developmental toxicity (fetal death) precludes derivation of 

an intermediate-duration oral MRL. Additional intermediate-duration dose-response studies determining 

NOAELs for the most sensitive end points, as well as the most sensitive species, would be useful for 

possible MRL derivation. Studies addressing interspecies differences could help to better characterize the 

relative sensitivity of monkeys and humans, particularly the possibility that monkeys are more sensitive 

than humans, as indicated by the high reproductive/developmental toxicity of PBBs in this species that 

has not been noted in PBB-exposed workers or the Michigan cohort. These studies could also help 

elucidate the toxicological significance of effects in the thyroid and other endocrine organs, particularly 

since the reproductive effects may be related to endocrine imbalance. 

Limited information is available on effects of PBBs in animals by inhalation or dermal exposure for 

intermediate durations (Millischer et al. 1980; Norris et al. 1975a; Waritz et al. 1977). Some inhalation 

data are available for octabromobiphenyl and decabromobiphenyl mixtures and some dermal data are 

available for octabromobiphenyl mixture, but intermediate-duration inhalation and dermal studies of 

FireMaster PBBs have not been performed. Studies of FireMaster FF-1 or FireMaster BP-6 would be 

particularly useful because these are likely to be the most toxic PBBs based on oral data and due to their 

higher content of potentially toxic congeners. Although the octabromobiphenyl mixture inhalation data 

are limited by numbers of animals, dose levels, and end points, and only one species (rat) was tested in 

the octabromobiphenyl mixture and decabromobiphenyl inhalation studies, it appears that these PBB 

mixtures are not highly toxic. Due to the inadequacies of the octabromobiphenyl mixture data and lack of 

any information on inhalation toxicity of the likely more potent FireMaster mixtures, there is insufficient 

basis for deriving an intermediate inhalation MRL. Although intermediate-duration inhalation studies of 

FireMaster PBBs would be particularly relevant to MRL derivation, they may not be practical due to the 
PBBs and PBDEs 323 

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low volatilization potential of PBBs. The intermediate-duration dermal studies of octabromobiphenyl 

mixtures revealed some skin irritation in rabbits but no sensitization in guinea pigs. Additional studies 

could corroborate the potential for dermal irritation by PBBs and are relevant because the skin is a route 

of concern for exposure at or near hazardous waste sites. Intermediate-duration inhalation and dermal 

exposure studies of PBB-contaminated soil (e.g., bioavailability studies) that identify thresholds would be 

especially useful for risk assessment at a hazardous waste site. 

Polybrominated Diphenyl Ethers. Available intermediate-duration oral studies in animals indicate that 

the liver and thyroid are the main systemic targets of repeated exposures to lower brominated PBDEs as 

shown by effects that mainly include enlargement and histological alterations in both organs and changes 

in serum levels of thyroid hormones, particularly decreases in serum T4 (Carlson 1980a; IRDC 1976, 

1977; WIL Research Laboratories 1984; Zhou et al. 2001, 2002). Hepatic effects occurred at the lowest 

LOAEL and were used as the basis for an intermediate oral MRL, but thyroid effects occurred at doses as 

nearly as low as those causing liver effects and the data are insufficient to clearly characterize liver effects 

as more critical than effects in the thyroid. Studies designed to identify NOAELs for octa-, and 

pentaBDE could provide a better basis for an intermediate duration MRL, as well as help to ascertain the 

most appropriate PBDE mixture on which to base a chronic MRL for lower brominated PBDEs. 

Essentially all of the available data on thyroid effects of lower brominated PBDEs have been obtained 

from oral studies in rats. It is speculated that the extent that PBDEs affect circulating levels of thyroid 

T4 or T3 might vary with species, and rats are often regarded as more sensitive than humans. As indicated 

in the Comparative Toxicokinetics section below, specific evidence is needed to determine whether 

PBDEs are likely to affect thyroid function in humans, or if humans are less sensitive to these effects than 

rats. Studies designed to elucidate the mechanism(s) of action for thyroid and other effects of lower 

brominated PBDEs would help to better understand how the animal toxicity data can best be used to 

identify target end points and assess health risks in humans. 

A commercial decaBDPE product (97.34% DBDPO, 2.66% nonaBDE and octaBDE) was administered to 

groups of 25 mated female Sprague-Dawley rats by gavage in corn oil in daily doses of 0, 100, 300, or 

1,000 mg/kg/day on gestation days 0 through 19 in a GLP-compliant study (Hardy et al. 2002). Each 

female was sacrificed on gestation day 20 and necropsied. End points included maternal clinical 

observations, maternal body weight/weight gain and food consumption, maternal gravid uterine and liver 

weights, maternal gross lesions, total number of corpora lutea, uterine implantations, early and late 

resorptions, viable and nonviable fetuses, and fetal weight and sex. Fetuses were examined grossly (all 

fetuses), evaluated for skeletal/cartilaginous malformations and ossification variations (approximately 
PBBs and PBDEs 324 

5. HEALTH EFFECTS 

half of each litter), and evaluated for visceral malformations (remaining fetuses). No treatment-related 

effects on any maternal or fetal endpoints were observed, indicating that 1,000 mg/kg/day was the 

NOAEL for maternal and developmental toxicity. 

One intermediate-duration oral toxicity study has been conducted for high purity decaBDE. In this study, 

multiple dose levels of a commercial decaBDE product (94B97% pure) was fed to rats and mice for 

13 weeks (NTP 1986). Comprehensive histological examinations were performed, but were limited to the 

control and high-dose groups, and no hematology, clinical chemistry, or urine indices, or thyroid hormone 

levels were evaluated. There were no compound-related clinical signs, deaths, body weight or food 

consumption changes, gross pathology, or histopathology, indicating that the highest tested doses in rats 

and mice (8,000 and 9,500 mg/kg/day, respectively) were intermediate-duration NOAELs for systemic 

toxicity. A developmental toxicity study in rats identified a NOAEL of 1,000 mg/kg/day (Hardy et al. 

2002). Because doses of decaBDE higher than 1,000 mg/kg/day have not been tested for developmental 

toxicity, and the NTP (1986) study indicates that this dose is also a NOAEL for systemic toxicity, the 

1,000 mg/kg/day developmental toxicity NOAEL was used to derive the MRL. Additional systemic and 

developmental toxicity studies could provide an indication of NOAEL/LOAEL thresholds and possibly 

provide a better basis for an intermediate-duration oral MRL for decaBDE. 

The inhalation database for intermediate-duration exposure to PBDEs consists of one well-conducted 

13-week unpublished industry study of octaBDE in rats (Great Lakes Chemical Corporation 2001). 

Hepatic, nasal, lung, thyroid, and ovarian effects were observed, and a NOAEL for changes in thyroid 

hormone levels was used as the basis for estimation of an intermediate-duration inhalation MRL. 

Chronic-Duration Exposure and Cancer. 

Polybrominated Biphenyls. Information on chronic systemic toxicity of PBBs in animals is limited to an 

oral bioassay showing hepatic, gastric, hematologic, and/or thyroid effects in rats and mice (NTP 1992), 

and a study showing effects on skin, stomach, and body weight in two monkeys (Allen et al. 1979; 

Lambrecht et al. 1978). Although limited by the number of studies and species, the available chronic 

animal data corroborate the results of intermediate-duration studies with respect to the observed effects. 

Additional studies would be necessary to determine the most sensitive animal target organ and species for 

chronic exposure and to provide a basis for an MRL, as serious hepatic changes as well as weight loss, 

decreased survival, and developmental effects occurred at the lowest tested dosages. Because PBBs are 

no longer being produced, exposure is most likely to occur at a contaminated waste site. Therefore, 

chronic studies examining the effects of PBB-contaminated soil following oral, inhalation, and dermal 
PBBs and PBDEs 325 

5. HEALTH EFFECTS 

exposure (i.e., bioavailability studies) would be particularly useful. Evaluations of the thyroid would be 

particularly informative because intermediate-duration animal studies indicate that the thyroid may be a 

particularly sensitive target organ. 

There is sufficient evidence that commercial hexabromobiphenyl mixtures (FireMaster FF-1) are 

hepatocarcinogenic in rats and mice following acute, intermediate, and/or chronic exposure (Groce and 

Kimbrough 1984; Kimbrough et al. 1978b; NTP 1983, 1992). Additional animal studies could provide 

useful information on interspecies differences and carcinogenesis of other PBB mixtures. 

Polybrominated Diphenyl Ethers. One chronic study of high purity decaBDE has been conducted. In 

this study, a commercial decaBDE product (94B97% pure) was fed to rats and mice for 103 weeks (NTP 

1986). Comprehensive gross and histological examinations were performed on all animals, but no 

hematology, clinical chemistry, or urine indices, or thyroid hormone levels, were evaluated. The lowest 

tested dose in the study, 1,120 mg/kg/day in male rats, was a LOAEL for a liver lesion (neoplastic 

nodules) that is precancerous and associated with thrombosis in the same tissue, precluding estimation of 

an MRL. Additional chronic dose-response information is needed to provide information on the 

NOAEL/LOAEL threshold and an appropriate basis for derivation of a chronic MRL for decaBDE. 

Neoplastic effects in this study included increased incidences of neoplastic nodules in the liver in the male 

and female rats and hepatocellular adenoma or carcinoma (combined) in the male mice. Slightly elevated 

incidences of thyroid gland follicular cell tumors were additionally observed in exposed male mice, 

although the increases were equivocal. 

Information on the chronic oral toxicity data of lower brominated PBDEs is available from a study in 

which rats were fed a 77.4% pure commercial decaBDE mixture (containing 21.8% nonaBDE and 0.8% 

octaBDE) for approximately 2 years (Kociba et al. 1975; Norris et al. 1975b). Evaluations that included 

clinical signs, body weight, food consumption, hematology, clinical chemistry, urine indices, and 

comprehensive histological examinations showed no exposure-related effects. The highest tested dose 

(1 mg/kg/day) was a NOAEL, but this effect level is not an appropriate basis for MRL estimation due to 

insufficient sensitivity of the study. In particular, a chronic oral MRL based on this study would be 

higher than the intermediate MRL. No exposure-related neoplastic changes were found, but the power of 

this study to detect carcinogenic effects is limited by the low dose levels. Considering the limitations of 

the available data, well-designed chronic toxicity studies of pentaBDE and/or octaBDE are needed to 

provide adequate bases for MRL derivation and cancer assessment for lower brominated PBDEs. 

Evaluations that include the thyroid and neurobehavioral end points would be particularly informative 
PBBs and PBDEs 326 

5. HEALTH EFFECTS 

because acute and intermediate-duration oral studies indicate that the thyroid and developing central 

nervous system are particularly sensitive targets for lower brominated PBDEs. 

Genotoxicity. 

Polybrominated Biphenyls. No information is available regarding potential genotoxic effects of PBBs in 

exposed humans. PBB mixtures or congeners were not genotoxic in any of the prokaryotic or eukaryotic 

animal systems tested. These include in vitro assays with S. typhimurium and E. coli bacteria (Haworth et 

al. 1983; Millischer et al. 1980; NTP 1983; Rossman et al. 1991), a host-mediated assay with 

S. typhimurium (Millischer et al. 1980), and in vitro assays with hamster cells (Galloway et al. 1987; 

Kavanagh et al. 1985; Williams et al. 1984), rat liver cells (Kavanagh et al. 1985; Williams et al. 1984), 

mouse liver and lymphoma cells (Myhr and Caspary 1991; Williams et al. 1984), and human fibroblasts 

(Williams et al. 1984). PBBs also were inactive in in vivo unscheduled DNA synthesis assays with rat 

and mouse hepatocytes (Mirsalis et al. 1985, 1989) and in a micronucleus test with mice (Millischer et al. 

1980). However, only some of these studies tested commercial PBB mixtures (Kavanagh et al. 1985; 

Millischer et al. 1980; Myhr and Caspary 1991; NTP 1983; Rossman et al. 1991; Williams et al. 1984). 

Additional studies of commercial mixtures could more fully characterize the genotoxic potential of PBBs, 

and provide information regarding differences in potencies of different mixtures and the sensitivities of 

different organisms. Cytogenic analysis of human populations exposed to PBBs in occupational settings, 

or exposed by consumption of food contaminated with PBBs, might make it possible to more adequately 

assess the genotoxic potential of these compounds in humans. 

Polybrominated Diphenyl Ethers. A limited amount of information has been published on the 

genotoxicity of PBDEs. Cytogenetic examination of bone marrow cells showed no increase in 

aberrations in maternal and neonatal rats following maternal oral exposure to a 77.4% decaBDE mixture 

(containing 21.8% nonaBDE and 0.8% octaBDE) for 90 days prior to mating and during mating, 

gestation, and lactation (Norris et al. 1973, 1975a). In vitro assays found that decaBDE did not induce 

gene mutations in bacteria (S. typhimurium) or mammalian cells (mouse lymphoma L5178Y cells), and 

did not induce sister chromatid exchange or chromosomal aberrations in Chinese hamster ovary cells 

(NTP 1986). In vitro exposure to the single congeners 2,2',4,4'-tetraBDE (BDE 47), 3,4-diBDE 

(BDE 12), and 2-monoBDE (BDE 1) caused increased recombinogenic activity in Chinese hamster SPD8 

and Sp5V79 cells (Helleday et al. 1999). Although the weight of available evidence indicates that 

decaBDE is not genotoxic, studies using lower brominated mixtures and a wider variety of assay types 

would help to better characterize the genotoxic potential of PBDEs. 
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Reproductive Toxicity. 

Polybrominated Biphenyls. A limited amount of information is available regarding reproductive effects 

in humans after exposure to PBBs. No evidence for PBBs-related effects on sperm counts, motility, or 

sperm morphology was found in a group of male Michigan workers exposed to PBBs by inhalation or 

dermal contact (Rosenman et al. 1979). No relationship was found between serum levels of PBBs and the 

frequency and duration of lactation in women exposed during the Michigan contamination episode 

(Thomas et al. 2001). 

Although no alterations in fertility or litter size were observed in mink fed PBB-containing diets prior to 

breeding and during pregnancy (Aulerich and Ringer 1979; Ringer et al. 1981) or in the F1 or F2

generations of female F0 rats fed PBB-containing diets during postimplantation gestation through weaning 

(McCormack et al. 1981), implantation was completely blocked in 40–67% of female rats exposed by 

gavage to PBBs between GDs 0 and 14 (Beaudoin 1979). Additionally, a lengthening of the menstrual 

cycle and prolonged implantation bleeding with decreased serum progesterone were observed in two of 

seven female monkeys fed a PBB-containing diet prior to and during pregnancy (Allen et al. 1979; 

Lambrecht et al. 1978). The dosage causing these reproductive effects in monkeys was the lowest tested 

in any intermediate-duration study of PBBs. In addition, alterations of male reproductive organs in rats 

(Gupta and Moore 1979) and in a monkey (Allen et al. 1978) have been observed after intermediateduration exposure to lethal oral doses of PBBs. Histopathological alterations were not observed in male 

or female reproductive organs after intermediate- or chronic-duration, oral exposure of rats or mice to 

nonlethal doses of PBBs (NTP 1983, 1992). The animal data suggest that PBBs may cause adverse 

effects on reproductive organs and their function(s) and that reproductive organ functions during the early 

phases of pregnancy may be particularly sensitive to PBBs. Additional studies in animals exposed by oral 

and other routes, including multi-generation studies with pre-breeding exposure to assess effects on 

fertility in both males and females, might help to further identify the reproductive processes affected by 

PBBs and to determine the dose-response relationships. Studies elucidating the NOAEL region and 

relative susceptibility of sensitive species (e.g., monkeys) to reproductive and developmental effects 

would be particularly useful, as these data could enable derivation of an intermediate oral MRL. 

Polybrominated Diphenyl Ethers. Information on the reproductive toxicity of PBDEs is limited to a 

single one-generation oral study of decaBDE in rats that found no exposure-related functional effects 

(Dow Chemical Co. 1975; Norris et al. 1975a, 1975b). Evaluation for histologic effects on reproductive 

organs in the available studies of PBDEs has generally reported no detectable effects. Tests of octaBDE 

and/or pentaBDE, particularly second-generation studies designed to assess effects on fertility in both 
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5. HEALTH EFFECTS 

sexes, would better characterize the reproductive toxic potential of PBDEs and assure the adequacy of the 

intermediate oral MRL. 

Developmental Toxicity. 

Polybrominated Biphenyls. No studies were located regarding developmental effects in humans or 

animals after inhalation or dermal exposure to PBBs. Studies of fetal mortality rates in Michigan 

(Humble and Speizer 1984) and of physical and neuropsychological development in Michigan children 

exposed in utero or in early infancy during the peak of the 1973 contamination episode (Schwartz and 

Rae 1983; Seagull 1983; Weil et al. 1981) did not conclusively correlate the ingestion of PBBs with 

developmental effects. Oral acute-, intermediate-, and chronic-duration studies of FireMaster FF-1 or 

FireMaster BP-6 in several species have reported fetotoxic and developmental effects, including 

embryolethality or increased mortality among nursing young (Allen et al. 1979; Beaudoin 1977, 1979; 

Groce and Kimbrough 1984; Lambrecht et al. 1978; Luster et al. 1980), fetal malformations (Beaudoin 

1977; Corbett et al. 1975; Waritz et al. 1977), growth retardation in offspring (Allen et al. 1979; Aulerich 

and Ringer 1979; Corbett et al. 1975; Groce and Kimbrough 1984; Lambrecht et al. 1978; McCormack et 

al. 1981; Meserve et al. 1992; Ringer et al. 1981), liver effects in offspring (Moore et al. 1978; Werner 

and Sleight 1981), and performance deficits in tests of operant behavior in offspring (Henck and Rech 

1986; Tilson 1992). A limited amount of data is available for octabromobiphenyl and 

decabromobiphenyl mixtures, which indicates that these PBBs are less developmentally toxic than 

FireMaster FF-1 or FireMaster BP-6 (Millischer et al. 1980; Waritz et al. 1977). Because FireMaster 

FF-1 caused developmental effects in monkeys at the lowest dosage tested in any study of PBBs, a 

chronic oral MRL could not be calculated; studies determining developmental NOAELs in sensitive 

species, therefore, would be particularly relevant. Additional studies regarding inhalation or dermal 

exposure to PBBs might help to determine whether or not the developmental toxicity of PBBs is routespecific. Studies on the mechanism(s) of action of PBBs in different animal species may provide a better 

understanding of the physiological and biochemical basis for the developmental toxicity of PBBs and a 

better basis for extrapolating from animal data in the evaluation of the hazard presented by PBBs to the 

development of human fetuses and children. 

Polybrominated Diphenyl Ethers. Oral developmental toxicity studies have shown no evidence of 

teratogenicity of deca-, octa-, and pentaBDE in rats and rabbits, although fetotoxic effects, including 

skeletal ossification variations at maternally toxic doses, have occurred with lower brominated mixtures 

(Argus Research Laboratories 1985b; Breslin et al. 1989; Dow Chemical Co. 1975, 1985; Hardy et al. 

2002; Life Science Research Israel Ltd. 1987; Norris et al. 1975a, 1975b; WIL Research Laboratories 
PBBs and PBDEs 329 

5. HEALTH EFFECTS 

1986). Effects of gestational exposure to lower brominated PBDEs included minimally increased postimplantation loss in rats and increased skeletal variations in rabbits at octaBDE doses as low as 10 and 

15 mg/kg/day, respectively (Argus Research Laboratories 1985b; Breslin et al. 1989; Life Science 

Research Israel Ltd. 1987). Exposure to decaBDE at level as high as 1,000 mg/kg/day (highest tested 

dose) caused no fetal or maternal effects in rats (Hardy et al. 2002). The available evidence appears to 

adequately show that teratogenicity and fetal toxicity is not a critical effect of concern for either lower 

brominated PBDEs or decaBDE. However, there is increasing evidence that the developing nervous 

system is a sensitive target of particular PBDE congeners, including decaBDE, as shown by impairments 

in tests of spontaneous motor behavior and learning and memory in adult mice exposed early in life 

(Branchi et al. 2001, 2002; Eriksson et al. 1999, 2001a, 2002a; Viberg et al. 2001a, 2001b, 2002a, 2002b, 

2003a) (see neurotoxicity data needs section). 

Immunotoxicity. 

Polybrominated Biphenyls. Information regarding immunological effects of PBBs in humans is 

equivocal. Some groups of investigators reported altered lymphocyte transformation responses in 

subjects accidentally exposed to PBBs through contaminated food (Bekesi et al. 1978, 1985; Roboz et al. 

1985). Other investigators could not confirm this in the same populations (Landrigan et al. 1979; Silva et 

al. 1979). Carefully designed follow-up studies of these populations would provide valuable information 

regarding possible immunological effects of PBBs. Additional research on the binding of PBBs with 

different plasma fractions could be fruitful, since it appears that on a per cell basis in exposed subjects, 

there is ≈100-fold excess of PBB in white cell fractions, compared to the erythrocyte fraction (Roboz et 

al. 1980, 1985). Acute oral data in rats and mice provided information regarding histopathology of the 

thymus, spleen, and lymph nodes (Fraker 1980; Fraker and Aust 1978; Gupta et al. 1981). Data from oral 

intermediate-duration studies in experimental animals suggest that the immune system may be one of the 

most sensitive targets for PBBs (Farber et al. 1978; Fraker 1980; Vos and van Genderen 1973, 1974). 

PBBs decreased the resistance of mice to infection by reducing antibody production (Loose et al. 1981), 

decreased the responsiveness of lymphocytes to mitogenic stimulation in rats and mice (Luster 1978, 

1980) and pigs (Howard et al. 1980), altered thymus weight in rats (NTP 1983), and caused thymus 

atrophy in dogs (Farber et al. 1978), guinea pigs (Vos and van Genderen 1973), and cattle (Moorhead et 

al. 1977). No studies were located regarding the immunological effects of PBBs in animals after 

inhalation or dermal exposure. Due to the relatively low vapor pressure of PBBs, inhalation is not a 

predominant route of exposure. Additional oral studies using a battery of immunological tests would be 

useful to further define the immunological effects of PBBs. 
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Polybrominated Diphenyl Ethers. Information regarding the immunosuppressive potential of PBDE 

mixtures is essentially limited to evidence from acute-duration oral studies in animals exposed to 

relatively high doses of pentaBDE. The plaque-forming splenic cell antibody response to injected sheep 

red blood cells was significantly reduced in mice exposed to 72 mg/kg/day pentaBDE for 14 days 

(Fowles et al. 1994), and in vitro production of IgG immunoglobulin from pokeweed mitogen-stimulated 

splenocytes was reduced in mice exposed to 36 mg/kg/day pentaBDE for 14 days (Darnerud and 

Thuvander 1998). Other 14-day studies in mice found no changes in natural killer cell activity at 

≤72 mg/kg/day (Fowles et al. 1994) or numbers of splenic and thymic lymphocyte subsets at 

≤36 mg/kg/day (Darnerud and Thuvander 1998), although 18 mg/kg/day of the single congener 

2,2’,4,4’-tetraBDE (BDE 47) caused significantly reduced numbers of total lymphocytes and CD4+, 

CD8+, and CD45R+ subtypes in the spleen (Darnerud and Thuvander 1998). In a 2-year study, NTP 

(1986) reported that exposure to 2,240 mg/kg/day of decaBDE in the diet resulted in a significant increase 

in splenic fibrosis; while not a measure of effects on immune function, it does indicate a treatment-related 

effect on an immune tissue. Additional oral studies using a battery of immunological tests and a lower 

range of doses would serve to better characterize the immunotoxic potential of PBDEs. 

Neurotoxicity. 

Polybrominated Biphenyls. One study was located regarding neurological effects in humans after 

inhalation and/or dermal exposure to PBBs (Brown et al. 1981). No studies were located regarding 

neurological effects in animals after inhalation or dermal exposure to PBBs. Although neurological 

symptoms were reported with some frequency by certain residents of Michigan who were likely to have 

consumed PBB-contaminated food, several studies of Michigan residents (including workers in PBB 

manufacturing who presumably were exposed predominately by inhalation and dermal contact) found no 

statistically significant associations between levels of PBBs in serum or fat (from oral or dermal exposure 

to PBBs) and frequencies of subjectively reported neurological symptoms or performance on neuropsychological tests (Anderson et al. 1978c, 1979; Barr 1980; Brown and Nixon 1979; Brown et al. 1981; 

Landrigan et al. 1979; Stross et al. 1981; Valciukas et al. 1978, 1979). Studies of the neuropsychological 

development of children exposed in utero or in early infancy, likewise, were inconclusive in establishing 

an association with PBB exposure (Schwartz and Rae 1983; Seagull 1983; Weil et al. 1981). Subtle 

effects in neurobehavioral tests were found in rodents, including decreased motor activity (Geller et al. 

1979; Tilson and Cabe 1979) and hind limb weakness (Cabe and Tilson 1978) after intermediate-duration, 

oral exposure and performance deficits in tests of learning behavior in the offspring of female mice and 

female rats exposed during gestation and lactation (Henck and Rech 1986; Henck et al. 1994; Tilson 

1992). Histopathological alterations of brain or spinal nerve tissue revealed no abnormalities in rats or 
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5. HEALTH EFFECTS 

mice after intermediate- or chronic-duration oral exposure (NTP 1983, 1992). Periodic neurobehavioral 

testing of animals exposed to PBBs at multiple doses for chronic durations would be useful for 

determining if longer-term exposure leads to more severe neurological effects than those observed with 

intermediate-duration exposures. 

Polybrominated Diphenyl Ethers. A limited amount of information is available on neurological effects 

of commercial PBDE mixtures. No clinical signs of neurotoxicity or neurohistopathology were observed 

in rats or mice exposed to commercial decaBDE in dietary doses as high as 16,000–19,000 mg/kg/day for 

14 days, 8,000–9,000 mg/kg/day for 13 weeks, or 2,550–7,780 mg/kg/day for 103 weeks (NTP 1986). 

Although the high doses and extended exposure durations provided opportunities for the induction and/or 

development of clinical signs, the study is limited by lack of testing for subtle behavioral changes and 

neurodevelopmental effects. A commercial pentaBDE mixture was evaluated for several behavioral end 

points in offspring of rats that were perinatally exposed to 1–100 mg/kg/day by gavage on GD 6 through 

PND 21 (MacPhail et al. 2003; Taylor et al. 2002, 2003). Evaluation of the offspring as adults showed no 

alterations in motor or sensory development as assessed by motor activity, habituation, and auditory 

startle response, although suggestive decreases in fear conditioning were observed. 

Neurobehavioral effects of individual PBDE congeners were evaluated in mice that were exposed during 

perinatal and/or early postnatal periods to 2,2',4,4'-tetraBDE (BDE 47), 2,2',4,4',5-pentaBDE (BDE 99), 

2,2',4,4',5,5’-hexaBDE (BDE 153), or 2,2',3,3’,4,4',5,5’,6,6’-decaBDE (BDE 209) (Branchi et al. 2001, 

2002; Eriksson et al. 1999, 2001a, 2002a; Viberg et al. 2001a, 2001b, 2002a, 2002b, 2003a). Most of 

these studies used similar single oral dose experimental designs and evaluated spontaneous motor 

behavior and swim maze performance at 2–6 months of age. The findings collectively indicate that the 

nervous system is a target of particular PBDE congeners during a defined critical phase of neonatal brain 

development, as shown by mild impairments in spontaneous motor behavior and learning and memory in 

older mice. One study used a different experimental design in which mice were exposed to BDE 99 from 

GD 6 to PND 21, and were evaluated using a variety of somatic (body weight gain, hair growth, day of 

eyelid and ear opening, day of incisor eruption) and neurobehavioral (righting reflex, forelimb stick 

grasping reflex, forelimb placing reflexes, negative geotaxis, screen grasping and climbing, pole grasping, 

ultrasonic vocalizations, homing test) end points during PNDs 2–22, as well as spontaneous activity 

endpoints on PNDs 22–120 (Branchi et al. 2001, 2002). Findings were suggestive of delayed 

sensorimotor development and altered spontaneous behavior. 
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Additional oral studies using comprehensive neurobehavioral test batteries are needed to better 

characterize the potential for PBDEs to cause neurotoxic effects in humans. Based on the limited 

available information (predominantly study abstracts), studies evaluating effects of exposure to lowerbrominated PBDEs and decaBDE commercial mixtures during critical postnatal periods of brain 

development would be particularly useful., 

Epidemiological and Human Dosimetry Studies. 

Polybrominated Biphenyls. Epidemiology studies of people exposed by ingesting PBB-contaminated 

food as a result of the 1973 Michigan PBB contamination episode or who were exposed occupationally in 

the manufacture or distribution of PBBs have not provided conclusive evidence that detectable effects 

have occurred as a result of exposure to PBBs (Anderson et al. 1978c, 1979; Barr 1980; Henderson et al. 

1995; Hoque et al. 1998; Humble and Speizer 1984; Landrigan et al. 1979; Thomas et al. 2001; Valciukas 

et al. 1978, 1979). Clinical examinations, including neuropsychological, liver function, and sperm count 

testing, of people who may have experienced the highest exposures did not conclusively identify 

particular effects or clinical signs associated with exposure (Brown and Nixon 1979; Brown et al. 1981; 

Rosenman et al. 1979; Schwartz and Rae 1983; Seagull 1983; Stross et al. 1981; Weil et al. 1981). No 

relationship was found between serum levels of PBBs and the frequency and duration of lactation in 

women exposed during the Michigan contamination episode (Thomas et al. 2001). A relationship 

between increasing serum levels of PBBs and increasing risk of breast cancer was indicated in casecontrol studies of women exposed during the Michigan episode (Henderson et al. 1995; Hoque et al. 

1998), but the results are only suggestive due to factors such as small number of cases, insufficient 

information on known breast cancer risk factors, and confounding exposures to other organochlorine 

chemicals. The evidence for an association between breast cancer and PBBs is inconclusive and warrants 

further study. Continued monitoring of the Michigan cohort for prevalence of other types of cancer as the 

cohort ages are also of interest, because lifetime and short-term exposure to PBBs are known to cause 

cancer in animals, and the residence time of PBBs in the body is expected to be long. If human exposure 

to PBBs is found to be occurring at a hazardous waste site, the nearby population should be studied for 

both exposure and effect data. 

Polybrominated Diphenyl Ethers. A limited amount of epidemiological information is available for 

PBDEs. Plasma levels of various organohalogen compounds, including the congener 2,2',4,4-tetraBDE, 

as well as serum hormone levels (free and total T3 and T4, TSH, free testosterone, follicle-stimulating 

hormone, lutenizing hormone, and prolactin), were analyzed in 110 men who consumed fatty fish from 

the Baltic Sea (Hagmar et al. 2001). There was a weak negative correlation between 2,2',4,4'-tetraBDE 
PBBs and PBDEs 333 

5. HEALTH EFFECTS 

(BDE 47) and plasma TSH after age adjustment, but BDE 47 could not explain more than 10% of the 

variance in TSH. No clear association was found between risk of non-Hodgkin’s lymphoma and adipose 

tissue levels of 2,2'4,4'-tetraBDE in a case-control study of 77 Swedish men and women (Hardell et al. 

1998; Lindstrom et al. 1998). 2,2'4,4'-TetraBDE was used as a marker for total PBDE exposure in both of 

these studies. No epidemiological data on PBDEs are available for non-oral exposure routes. 

Animal data raise particular concern for effects of PBDEs on the thyroid as well as possible 

immunological, neurodevelopmental, and carcinogenic effects of exposure. Epidemiological 

investigations are needed to better characterize the potential for PBDEs, both the lower PBDEs and 

decaBDE, to induce these and other kinds of effects as well as the relationship between PDBE body 

burden and the observed effects. Considering the possibility that PBDEs can be transferred to the fetus 

across the placenta and that greater amounts might be transferred to nursing infants via breast milk, as 

well as evidence that perinatal exposure to PCBs and other similar chemicals may induce subtle 

neurological damage and immunological and thyroid effects in children, transgenerational studies would 

be particularly informative. Limitations that are likely to constrain the epidemiological investigations, 

such as unmeasured PBDE exposure concentrations and lack of controls for confounding co-exposures, 

should be addressed. 

Biomarkers of Exposure and Effect. 

Exposure. 

Polybrominated Biphenyls. PBBs are stored primarily in adipose tissue and are present in serum and 

human milk of exposed populations. Several studies have shown that serum and adipose PBB levels are 

biomarkers of exposure (Blanck et al. 2000b; Brilliant et al. 1978; Humphrey and Hayner 1975; Lambert 

et al. 1990; Landrigan et al. 1979; Rosen et al. 1995; Wolff et al. 1982). It has been proposed that 

measurement of PBB levels in adipose tissue may be a more reliable prediction of body burden than 

serum levels because of the high adipose/serum PBB partition ratio (Anderson 1985). However, once a 

stable correlation between adipose/serum levels has been characterized, serum levels are a better choice 

for surveillance and monitoring (Anderson 1985). Further studies on the predictive value of levels of 

PBB (particularly congeners) in serum and adipose tissue in individuals exposed to PBBs for acute, 

intermediate, and chronic durations would provide valuable information that could lead to early detection 

of PBB exposure. 
PBBs and PBDEs 334 

5. HEALTH EFFECTS 

A potential biomarker of exposure to PBBs is related to their effect on the thyroid gland. Effects in 

exposed workers included increased serum thyrotropin, low or borderline low serum T4, and increased 

thyroid antimicrosomal antibody titers (Bahn et al. 1980), and effects in exposed rats included reduced 

levels of serum T4 and T3 (Akoso et al. 1982b; Allen-Rowlands et al. 1981; Byrne et al. 1987; Gupta and 

Moore 1979; Kasza et al. 1978a; Meserve et al. 1992; Norris et al. 1975a, 1975b; NTP 1983; Sepkovic 

and Byrne 1984; Sleight et al. 1978; Werner and Sleight 1981). Additional studies could better 

characterize thyroid effects of PBBs in humans and develop specific correlations between levels and 

duration of exposure and alterations in serum hormone levels. 

Polybrominated Diphenyl Ethers. PBDEs also accumulate in adipose tissue, serum, and breast milk of 

the general population due to their lipophilic characteristics. Concentrations of PBDEs in breast milk are 

useful, non-invasive markers of maternal body burdens and of in utero and lactational exposures, but 

body burden assessments are limited by a lack of time-trend data for PBDEs in the milk of U.S. 

populations (Hooper and McDonald 2000). Breast milk monitoring programs are needed to provide timetrend data and to verify findings that PBDE levels have been exponentially increasing in breast milk 

during the past 25 years (Norén and Meironyté 1998, 2000). Studies on the predictive value of levels of 

PBDEs in serum and adipose tissue could provide useful information for detection and monitoring of 

exposure. It should be noted, however, that solubilities in adipose and breast milk are likely to vary with 

the congener; for example, decaBDE is much less soluble in adipose than pentaBDE. These differences 

must be kept carefully in mind when designing studies evaluating PBDE exposure. 

A potential biomarker of exposure to PBDEs relates to their effect on the thyroid gland. Thyroid changes 

in rats and mice include reduced serum T4 levels with no changes in serum TSH (Darnerud and Sinjari 

1996; Fowles et al. 1994; Hallgren and Darnerud 1998; WIL Research Laboratories 1984; Zhou et al. 

2001, 2002). However, using thyroid changes as a biomarker may not be reliable, as thyroid changes are 

not specific to exposure to PBDEs and the effects associated with the thyroid in non-clinical studies are 

likely specific to the rodent and may or may not be directly relevant to the human. Additional studies 

could characterize thyroid effects of PBDEs in humans and develop specific correlations between levels 

and duration of exposure and alterations in serum levels of T4. 

Effect. 

Polybrominated Biphenyls. There are no specific biomarkers of effects for PBBs. Numerous studies 

have attempted to correlate serum and adipose PBB levels with an array of symptoms and health 
PBBs and PBDEs 335 

5. HEALTH EFFECTS 

complaints in PBB exposed subjects from the state of Michigan (Anderson et al. 1978a, 1978b, 1978c, 

1979; Bekesi et al. 1978; Humphrey and Hayner 1975; Landrigan et al. 1979; Stross et al. 1979). Thus 

far, no significant correlation has been found. Continued follow-up studies on the Michigan cohort would 

provide information on effects that may have a long latency, such as cancer. Elevated levels of two 

cytochrome P-450I-dependent enzymes were observed among PBB exposed subjects, relative to controls 

(Lambert et al. 1990). The thyroid is a sensitive target for PBBs and characteristic changes include 

reduced serum levels of T4 and other thyroid hormones (Akoso et al. 1982b; Allen-Rowlands et al. 1981; 

Byrne et al. 1987; Gupta and Moore 1979; Kasza et al. 1978a; Meserve et al. 1992; Norris et al. 1975a, 

1975b; NTP 1983; Sepkovic and Byrne 1984; Sleight et al. 1978; Werner and Sleight 1981), indicating 

that they are potential biomarkers of effect. Levels of CYP enzymes and thyroid hormones, however, are 

not specific for PBB exposure. Further studies designed to identify specific biomarkers of effects of 

PBBs would facilitate medical surveillance leading to early detection of potentially adverse health effects 

and possible treatment. Congener-specific analysis may be useful for characterizing dioxin-like health 

effects. 

Polybrominated Diphenyl Ethers. Biomarkers that could be used to characterize health effects caused 

specifically by PBDEs have not been identified. The thyroid is a critical target for PBDEs in animals, as 

discussed in Section 3.2.2.2, Endocrine Effects, and serum T4 is a potential biomarker of effect for these 

chemicals in humans. Although this biomarker is not specific to PBDEs because other antithyroid agents 

can have similar effects, changes in T4 could be considered to indicate potential impairment of health. 

Absorption, Distribution, Metabolism, and Excretion. 

Polybrominated Biphenyls. There are no quantitative data regarding absorption in humans via the 

inhalation route, but data from occupationally exposed individuals and individuals who ingested food 

contaminated with PBBs suggest that exposure by the oral or dermal route may lead to considerable 

accumulation of PBBs in tissues (Anderson et al. 1978c; Eyster et al. 1983; Landrigan et al. 1979). The 

animal data indicate that the main component of a commercial PBB mixture (2,2,'4,4',5,5'-hexabromobiphenyl) is efficiently absorbed by the oral route (Matthews et al. 1977; Tuey and Matthews 

1980). Data regarding absorption after inhalation exposure was limited to a single study (Waritz et al. 

1977). There are no data regarding absorption via the dermal route. No studies were located in which 

several doses of different PBB congeners were administered by the inhalation, oral, and dermal routes, 

and for various exposure periods. Such studies could provide information on the relationship between 

bromination patterns and absorption efficiency and rates of absorption by the different routes of exposure. 
PBBs and PBDEs 336 

5. HEALTH EFFECTS 

In addition, studies with different PBB mixtures could help determine possible interaction effects among 

congeners that could affect absorption. 

Distribution data are limited to qualitative information derived from cases of accidental ingestion of food 

contaminated with PBBs, cases of occupational exposure via dermal contact (Eyster et al. 1983; 

Landrigan et al. 1979) and autopsy reports (Miceli et al. 1985). These data suggest that PBBs distribute 

preferentially to tissues with high fat content regardless of the route of exposure. Data derived from oral 

administration of PBBs to animals indicate that PBBs are distributed first to liver and muscle and then to 

adipose tissue where they are stored (Domino et al. 1982; Lee et al. 1975; Matthews et al. 1977). Little 

information regarding distribution of PBBs could be drawn from the limited number of studies in animals 

administered PBBs by the inhalation or dermal routes. Additional well-conducted studies by these routes 

of exposure would provide useful information regarding possible route-dependent distribution patterns. 

Studies regarding distribution through the placenta after inhalation and dermal exposure were not 

available. 

Data regarding biotransformation of PBBs in humans are limited to individuals who accidentally 

consumed food contaminated with PBBs or who were exposed to PBBs in the workplace (Wolff and 

Aubrey 1978; Wolff et al. 1979a). The use of human cell systems in culture might be considered a useful 

addition to whole animal studies for studying the metabolic fate of PBBs. There are studies regarding the 

metabolism of some PBB congeners after oral administration to rats (Sparling et al. 1980), rabbits (Kohli 

et al. 1978), and pigs (Kohli and Safe 1976). However, the PBBs mainly studied were monobromobiphenyls and dibromobiphenyls, which are only trace components of FireMaster mixtures. Therefore, 

studies on the in vivo metabolism of the main components of commercial PBB mixtures would provide 

valuable information regarding the metabolic disposition of highly brominated congeners. A limited 

amount of information is available on the metabolism of PBBs in farm animals (e.g., dairy cows, 

chickens). This is a data gap because people exposed to PBBs during the Michigan PBB contamination 

episode were predominately exposed by consuming products of farm animals. Although information 

regarding metabolism after inhalation or dermal exposure is lacking, there is no evidence to suggest that 

other pathways would operate after exposure by these routes. 

Studies regarding urinary or fecal excretion of PBBs in humans were not located; however, elimination of 

PBBs through maternal milk is well documented (Brilliant et al. 1978; Eyster et al. 1983; Jacobson et al. 

1984; Landrigan et al. 1979). Fecal excretion of unabsorbed PBBs appears to be the main route of 

elimination of highly brominated congeners after oral exposure (Matthews et al. 1977; Norris et al. 1975a; 
PBBs and PBDEs 337 

5. HEALTH EFFECTS 

Rozman et al. 1982), whereas polar derivatives formed by lower brominated congeners appeared to be 

excreted mainly in the urine (Kohli and Safe 1976; Kohli et al. 1978; Sparling et al. 1980). Although data 

regarding excretion in animals after inhalation and dermal exposure were not located, there is no reason to 

believe that results from additional studies would reveal different patterns of excretion. 

Polybrominated Diphenyl Ethers. A limited amount of data is available on the toxicokinetics of PBDEs. 

There are data gaps in a number of areas, particularly for octaBDE and pentaBDE commercial mixtures 

and the tetra and hexa congeners that are most prevalent in the environment. Quantitative absorption 

studies could corroborate the conclusions on oral uptake in animals that are based on elimination and 

excretion data. Metabolism studies would help to characterize the enzymes involved as well as the 

transformation of some congeners to biologically active hydroxylated BDEs. There are still data gaps in 

the toxicokinetics of decaBDE, including an incomplete understanding of the debromination of decaBDE 

to lower brominated BDEs. 

Comparative Toxicokinetics. 

Polybrominated Biphenyls. The data suggest that there are qualitative differences in the toxicokinetic 

disposition of PBBs among humans and among animal species (Wolff and Aubrey 1978; Wolff et al. 

1979a). However, these differences appear to be highly dependent on the specific congener or mixture 

studied. In general, all species absorb PBBs, with varying efficiency, and accumulate PBBs in tissues 

rich in fat. Once absorbed, PBBs are distributed in a biphasic manner in all examined animal species 

(Domino et al. 1982; Ecobichon et al. 1983; Matthews et al. 1977). No studies were located that provide 

information regarding differences or similarities in metabolic disposition of PBBs between humans and 

animals. Limited data in humans indicate that fecal excretion of PBB residues occur (Eyster et al. 1983). 

Experimental data in animals suggest that the rate and extent of PBB elimination in the urine and feces 

are dependent on the degree and pattern of bromination (Kohli et al. 1978; Matthews et al. 1977; Sparling 

et al. 1980). Analysis of the excreta of humans exposed in the workplace and near hazardous waste sites 

would provide information regarding biotransformation and elimination kinetics in humans. In addition, 

similar target organs have been identified across animal species, but the database is not complete enough 

for identifying a most sensitive species. Although the toxicological data in humans are limited and 

inconclusive, adverse immune effects observed in humans (Bekesi et al. 1978) have also been observed in 

rats, mice, and pigs (Howard et al. 1980; Luster et al. 1978, 1980) suggesting that any of these species 

may represent a suitable animal model for humans. The only reported PBPK model for PBBs describes 

the distribution and body burden of the major component of FireMaster mixtures, 2,2',4,4',5,5'-hexabromobiphenyl (BB 153), in the rat (Tuey and Matthews 1980). The serum mean half-life of 6.5 years 
PBBs and PBDEs 338 

5. HEALTH EFFECTS 

predicted using this model is shorter than half-life values of approximately 12–29 years estimated using 

human sera data from the Michigan PBB cohort (Blanck et al. 2000b; Lambert et al. 1990; Rosen et al. 

1995); one possible reason for this difference is differences in body fat between humans and laboratory 

rodents resulting in a different distribution of the administered compound. This indicates a need for an 

improved PBPK model for extrapolating animal data to humans and/or for studies designed to produce 

data for improving the performance of PBPK analyses. 

Polybrominated Diphenyl Ethers. Insufficient data are available to determine whether qualitative 

differences in the toxicokinetic disposition of PBDEs exist between humans and animals and among 

animal species. Differences are likely to be dependent on the specific congener or mixture studied, and 

pharmacokinetic modeling studies could help to determine the validity of extrapolating data. Most of the 

available toxicokinetic studies of PBDEs have been performed in rats, and studies in other species could 

help to ascertain the most relevant animal model. 

The extent that PBDEs affect circulating levels of thyroid T4 or T3 might vary with species and rats are 

often regarded as more sensitive than humans. The main basis for this opinion seems to be studies 

showing that PBDEs affect binding of thyroid hormones to transthyretin (TTR), the primary transport 

protein in rats. Because TTR is not the major transport protein in humans, the findings have been 

interpreted as evidence that humans will be less sensitive than rats to thryoid effects of PBDEs. As 

discussed in Section 5.5.3, the greater sensitivity of rats is thought to be related to a smaller and more 

rapid turnover of the hormone pool in the rat thyroid, and to a more rapid clearance of secreted hormone 

in the rat; the latter being, in part, related to the absence of thyroid binding globulin (TBG) in rats. 

Whereas TTR is the major thyroid hormone binding protein in rats, TBG is the main binding protein in 

humans and most other mammals. Specific evidence is needed to determine whether PBDEs are likely to 

affect thyroid function in humans, or if humans are less sensitive to these effects than rats. 

Methods for Reducing Toxic Effects. 

Polybrominated Biphenyls. The mechanism by which PBBs enter the blood stream in humans is not 

known; consequently, there are no established methods for reducing absorption. Studies in experimental 

animals that could identify substances that prevent or delay absorption and that do not represent a toxic 

risk per se would be valuable. There are no established methods for reducing body burden in humans, but 

studies in animals and model simulations in humans indicate that reducing body fat markedly increases 

elimination of PBBs (Domino et al. 1982; Tuey and Matthews 1980). The effect of reduction of body fat 

(e.g., by dieting and exercising) in PBB-exposed humans has not been fully researched. 
PBBs and PBDEs 339 

5. HEALTH EFFECTS 

The mechanism of toxic action of PBBs is not completely understood and no methods exist to block the 

toxic response due to exposure to PBBs. A more complete characterization of the cytosolic AhR protein, 

to which some PBB congeners are thought to bind, and understanding of physiological effects of receptor 

blockage would be useful for the possible identification of blockers of AhR-mediated toxic effects. 

Further studies aimed at elucidating the nonreceptor-mediated mechanism of action of some PBBs would 

also be valuable. 

Polybrominated Diphenyl Ethers. The mechanism by which PBDEs enter the blood stream is not 

known, there are no established methods for reducing body burden of PBDEs, and the mechanisms of 

toxic action of PBDEs are incompletely understood. Types of studies that could address these data gaps 

and possibly provide information on reducing toxic effects of PBDEs are discussed in the preceding 

subsection on PBBs. 

Children’s Susceptibility. 

Polybrominated Biphenyls. Information on health effects of PBBs in children is available from several 

studies of the Michigan feed contamination episode. A 1976 study of Michigan children likely to have 

been exposed to PBBs found that the number of subjectively reported symptoms of ill health, including 

several symptoms of neurological effects, did not increase with increasing serum PBB levels, but rather 

decreased; general neurological examinations did not reveal a pattern of abnormality among the Michigan 

children (Barr 1980). Studies of fetal mortality rates in Michigan (Humble and Speizer 1984) and of 

physical and neuropsychological development in Michigan children exposed during the contamination 

episode (Schwartz and Rae 1983; Seagull 1983; Weil et al. 1981) did not conclusively correlate the 

ingestion of PBBs with effects. Neurobehavioral alterations have been observed in rats following 

gestational and lactational exposure to PBBs (Henck and Rech 1986; Henck et al. 1994; Tilson 1992). 

Other effects in offspring of rats exposed to PBBs during gestation and lactation include decreased serum 

levels of thyroid hormone levels (Meserve et al. 1992; Werner and Sleight 1981). These effects in 

offspring are consistent with evidence that the thyroid gland is an unequivocal target of PBBs in adult 

animals. No information was located on possible immunological effects of PBBs in children, and data in 

adult humans are limited and largely inconclusive (Bekesi et al. 1978; Landrigan et al. 1979; Roboz et al. 

1985; Silva et al. 1979; Stross et al. 1981), but exposure to PBBs caused altered immune responses in a 

variety of animal species, which suggests that children may also be affected. Continued assessment of 

children exposed to PBBs during the Michigan contamination episode, with particular emphasis on 
PBBs and PBDEs 340 

5. HEALTH EFFECTS 

evaluation of cognitive abilities, thyroid function, and immune competence, would help to better assess 

the susceptibility of children to PBBs. 

Polybrominated Diphenyl Ethers. No information was located on health effects of PBDEs in newborn or 

older children. Thyroid and neurobehavioral alterations have been observed in animals following preand/or postnatal exposure to commercial PBDE mixtures and single PBDE congeners, indicating that 

these are potential effects of concern in exposed children. Serum levels of thyroid T4 and T3 hormones 

were reduced in offspring of rats that were orally exposed to pentaBDE during gestation and lactation and 

in rats exposed as weanlings (Zhou et al. 2002). Alterations in spontaneous locomotion behavior and 

learning and memory ability were observed in mice that were tested at 2 months of age and as adults 

(4 months) following neonatal exposure to single low oral doses of the congeners 2,2',4,4'-tetraBDE 

(BDE 47) and 2,2',4,4',5-pentaBDE (BDE 99) (Eriksson et al. 1998, 1999, 2001a, 2002a). Effects were 

observed in adult mice treated at either 3 or 10 days of age, but not at 19 days of age, suggesting that there 

was a critical window for the induction of behavioral disturbances (Eriksson et al. 1999, 2002a). 

Additional studies are needed to better characterize the potential susceptibility of children to the effects of 

PBDEs on the thyroid and neurodevelopment, particularly considering the possibility that these effects are 

related to the dependence of central nervous system development on thyroid hormones. No information is 

available regarding the immunosuppressive potential of PBDEs in children or young animals, indicating 

that studies of immune competence in developing animals would also help to more fully assess children’s 

susceptibility to PBDEs. 

Child health data needs relating to exposure are discussed in Section 8.8.1 Identification of Data Needs: 

Exposures of Children. 

5.12.3 Ongoing Studies 

Ongoing studies that are relevant to health effects of PBBs and PBDEs, as identified in the Federal 

Research in Progress database (FEDRIP 2002) and the websites of various U.S. government agencies, are 

listed in Table 5-9. 

The Great Lakes Chemical Corporation is collaboratively working with Health Canada to assess potential 

effects of pentaBDE using a three-tier study approach (Biesemeier 2004). The first tier study, a 28-day 

repeated oral (gavage) toxicity study in rats, has been completed, but the draft report was not available as 

of June 2004. The basis of this study was mainly to serve in selecting gavage dose levels for the second 
PBBs and PBDEs 341 

5. HEALTH EFFECTS 

Table 5-9. Ongoing Studies on the Health Effects of PBBs and PBDEs 

Investigator Affiliation Research description Sponsor Source 

Anderson HA Wisconsin Department 

of Health and Family 

Services, Madison, 

Wisconsin 

To characterize human exposures to 

PCBs, PBDEs, and DDT from Great 

Lakes fish consumption and to 

determine mechanisms by which 

PBDEs in Great Lakes fish may act 

separately or synergistically with PCB 

exposure to impair human thyroid 

function 

EPA EPA 

2004d 

Hammock B University of California, 

Davis, California 

Support for testing hypotheses 

regarding the association of PBBs, 

PBDEs, and other known xenobiotic 

immunotoxicants and neurotoxicants 

with autism 

NIEHS/EPAa FEDRIP 

2002 

Hites RA and 

Bigsby M 

Indiana University, 

Bloomington, Indiana 

Compare the concentration of PBDEs 

in infants and analyze sediments and 

fish in the Great Lakes 

EPA EPA 

2004d 

Huwe JK Agricultural Research 

Service, Fargo, North 

Dakota 

Development of effective remediation 

procedures for PBDEs and other 

persistent organic pollutants in animal 

tissues and their environment 

USDA FEDRIP 

2003 

Karmaus W 

et al. 

Michigan State 

University, East Lansing, 

Michigan 

An effort to determine if exposure to 

halogenated organic compounds 

(including PBBs and PBDEs) via 

breastfeeding creates a risk to the 

immune system of the child 

EPA EPA 

2004d 

Ludewig G University of Kentucky, 

Lexington, Kentucky 

A multispecies approach to analyze 

the toxic effects of PBDEs on organs 

and development 

EPA FEDRIP 

2002 

Marcus M Emory University, 

Atlanta, Georgia 

Investigation of the effect PBBs have 

on pubertal development, reproductive 

health, and ovarian function 

NIEHS FEDRIP 

2003 

Omann GM Department of Veterans 

Affairs, Washington, DC 

The immunotoxicological potential of 

PBDEs found in the environment will 

be assessed using fish immune cells 

in an in vitro model 

Department 

of Veteran 

Affairs 

FEDRIP 

2003 

Palmer BD University of Kentucky 

Medical Center, 

Lexington, Kentucky 

Multidisciplinary/multispecies 

investigation of end points and 

mechanisms of action for PBDEs and 

PCDEs, including structure-activity 

relationships for endocrine disruption 

NIEHS/EPAa

 FEDRIP 

2002 

Raymer JH 

et al. 

Research Triangle 

Institute, Research 

Triangle Park, North 

Carolina 

Development of a PBPK animal model 

for PBDEs to estimate fetal exposure 

in humans and to determine the 

validity of the model by creating/using 

new analytical methods using cord 

blood and meconium 

EPA EPA 

2004d 
PBBs and PBDEs 342 

5. HEALTH EFFECTS 

Table 5-9. Ongoing Studies on the Health Effects of PBBs and PBDEs 

Investigator Affiliation Research description Sponsor Source 

Robertson L University of Kentucky 

Medical Center, 

Lexington, Kentucky 

Laboratory synthesis of pure PBDE 

and PCDE congeners and their 

metabolites 

NIEHS/EPAa FEDRIP 

2002 

Sikka HC State University of New 

York College at Buffalo, 

Buffalo, New York 

Disposition and metabolism of PBDEs 

in fish 

NOAAb FEDRIP 

2002 

Trosko JE Michigan State 

University, East Lansing, 

Michigan 

Epigenic effects of PBBs and other 

environmental toxicants on cellular 

communication pathways 

NIEHS/EPAa

 FEDRIP 

2002 

Vos J Bilthoven, Netherlands Risk assessment of brominated flame 

retardants for human health and 

wildlife 

EU EU 2004 

Willett LB Ohio State Universities, 

Wooster, Ohio 

Develop methods to monitor the 

occurrence of PBBs and other 

xenobiotics in the environment of 

cattle; create methods that will reduce 

or eliminate exposure of cattle and the 

food products they produce to these 

xenobiotics; determine mechanisms 

by which xenobiotics are transported, 

bound, and mobilized in vivo; describe 

the pharmacokinetics; and to study 

target organ modifications in cattle 

caused by xenobiotic chemicals that 

result in cellular or metabolic 

alterations 

Department 

of Agriculture 

FEDRIP 

2003 

a

NIEHS/EPA Superfund Basic Research Program 

b

National Sea Grant College Program 

DDT = dichlorodiphenyltrichloroethane; EPA = U.S. Environmental Protection Agency; EU = European Union; 

NIEHS = National Institute of Environmental Health Sciences; NOAA = National Oceanic and Atmospheric 

Administration; PCB = polychlorinated biphenyls; PBPK = physiologically based pharmacokinetic; USDA = U.S. 

Department of Agriculture 
PBBs and PBDEs 343 

5. HEALTH EFFECTS 

tier study, a one-generation reproductive study in rats, of which the draft protocol is currently being 

written (as of June 2004). This protocol is being drafted with a pharmacokinetic segment, but it has not 

yet been confirmed whether it will be retained in the final study design. The results of the one-generation 

study will be used to select gavage dose levels for the third tier study, a two-generation reproductive study 

in rats with endocrine, developmental neurobehavioral, and other end point segments. The entire program 

is not expected to be completed until near the end of 2005. 

NTP is in the process of designing a study to assess the toxicity of tetraBDE (NTP 2004). 

PBBs and PBDEs 345 

6. CHEMICAL AND PHYSICAL INFORMATION 

6.1 CHEMICAL IDENTITY 

Polybrominated Biphenyls. PBBs are a class of structurally similar brominated hydrocarbons in which 

2–10 bromine atoms are attached to the biphenyl molecule. Monobrominated structures (i.e., one 

bromine atom attached to the molecule) are often included when describing PBBs. The general chemical 

structure of PBBs is shown below: 

1'

2' 3'

4'

6' 5'

1

3 2

4

5 6 Br Brm n

where m + n = 1–10 

It can be seen from the structure that a large number of brominated compounds are possible. The 

209 possible compounds for PBBs are called “congeners”. However, the number of PBB congeners that 

actually exist in commercial PBB mixtures is much less compared to polychlorinated biphenyls (PCBs). 

Typically, only a subset of the 209 possible congeners is observed for PBBs. PBBs can be categorized by 

degree of bromination. The term “homolog” is used to refer to all PBBs with the same number of 

bromines (e.g., tribromobiphenyls). Based on the number of bromine substituents, there are 

10 homologous groups of PBBs (monobrominated through decabrominated). Each homologous group 

contains one or more congeners. The mono-, di-, tri-, tetra-, penta-, hexa-, hepta-, octa-, nona-, and 

decabromo congeners can exist in 3, 12, 24, 42, 46, 42, 24, 12, 3, and 1 forms, respectively. Homologs 

with different substitution patterns are referred to as isomers. For example, the group of dibromobiphenyl 

homologs contains 12 isomers. The numbering system for PBBs is also shown above. Positions 2, 2', 6, 

and 6' are called ortho positions, positions 3, 3', 5, and 5' are called meta positions, and positions 4 and 

4' are called para positions. In a PBB molecule, the benzene rings can rotate around the bond connecting 

them; the two extreme configurations are planar (the two benzene rings are in the same plane; dihedral 

angle=0°) and nonplanar (the two benzene rings are in perpendicular planes to each other; dihedral 

angle=90°). The degree of planarity is largely determined by the number of substitutions in the ortho

positions. The replacement of hydrogen atoms in the ortho positions with larger bromine atoms forces 

the benzene rings to adopt a configuration with a larger dihedral angle or a nonplanar configuration. The 

benzene rings of non-ortho substituted PBBs, as well as mono-ortho substituted PBBs, may assume a 

small dihedral angle (in which the dihedral angle is small, but >0°) or “near” planar configuration. These 
PBBs and PBDEs 346 

6. CHEMICAL AND PHYSICAL INFORMATION 

molecules are referred to as planar or coplanar congeners. The benzene rings of other congeners cannot 

assume a planar or coplanar configuration and are referred to as nonplanar congeners (Hardy 2002). 

Like PCBs, the 209 congeners for PBBs are arranged in ascending numerical order using a numbering 

system developed by Ballschmiter and Zell (1980) that follows the IUPAC rules of substituent 

characterization of biphenyls. The resulting numbers assigned by Ballschmiter and Zell (which are also 

referred to as congener, IUPAC, or BZ numbers) are widely used for identifying individual congeners of 

PBBs. For example, the PBB congener, 2,2',4,4',5,5'-hexabromobiphenyl, may be referred to as BB 153 

in this document. The identities of several PBB congeners are shown in Table 6-1 (WHO 1994a, 1994b). 

Michigan Chemical Corporation, the major producer of PBBs from 1970 to 1976, marketed mixtures of 

PBBs under the trade name FireMaster (e.g., BP-6 and FF-1). However, the FireMaster trade name has 

also been used for other brominated flame retardants using different numerical designations. Other 

former producers of PBBs in the United States included White Chemical Corporation (Bayonne, New 

Jersey) and Hexcel Corporation (Sayreville, New Jersey), which both produced technical mixtures of 

octabromobiphenyl and decabromobiphenyl until 1979. The trade names of some commercial PBB 

mixtures formerly produced in other countries are: Berk Corporation, Great Britain (e.g., BerkFlam, 

Flammex); Chemische Fabrik Kalk, Germany (e.g., Bromkal); and Ugine Kuhlmann (now Atofina in 

France) (e.g., Adine). 

The chemical identities of hexabromobiphenyl, octabromobiphenyl, decabromobiphenyl (BB 209), and 

BB 153, the most abundant congener in commercial FireMaster FF-1 and FireMaster BP-6, are listed in 

Table 6-2. 

Polybrominated Diphenyl Ethers. PBDEs are a class of structurally similar brominated hydrocarbons, in 

which 2–10 bromine atoms are attached to the diphenyl ether molecule. Monobrominated structures (i.e., 

one bromine atom attached to the molecule) are often included when describing PBDEs. The general 

chemical structure of PBDEs is shown below: 

O

Brn Brm

1'

2'

3'

4'

5'

6'

1

2

3

4

5

6

where m + n = 1–10 
PBBs and PBDEs 347 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-1. Chemical Identity of Polybrominated Biphenyl (PBB) and 

Polybrominated Diphenyl Ether (PBDE) Congeners 

CAS No.b IUPAC 

No.a

Compound/ 

substituents Brominated biphenyls (BB)c Brominated diphenyl ethers (BDE)d

 Biphenyl 92-52-4 92-52-4 

 Monobromo- 26264-10-8 101-55-3 

1 2 2052-07-7 

2 3 2113-57-7 

3 4 92-66-0 

 Dibromo- 27479-65-8 2050-47-7 

4 2,2’ 13029-09-9 

5 2,3 115245-06-2 

6 2,3’ 49602-90-6 

7 2,4 53592-10-2 

8 2,4’ 49602-91-7 

9 2,5 57422-77-2 

10 2,6 59080-32-9 

11 3,3’ 16400-51-4 

12 3,4 60108-72-7 

13 3,4’ 57186-90-0 

14 3,5 16372-96-6 

15 4,4’ 92-86-4 

 Tribromobiphenyl 51202-79-0 49690-94-0 

16 2,2’,3 

17 2,2’,4 

18 2,2’,5 59080-34-1 

19 2,2’,6 

20 2,3,3’ 

21 2,3,4 

22 2,3,4’ 

23 2,3,5 

24 2,3,6 

25 2,3’,4 

26 2,3’,5 59080-35-2 

27 2,3’,6 

28 2,4,4’ 6430-90-6 

29 2,4,5 115245-07-3 

30 2,4,6 59080-33-0 

31 2,4’,5 59080-36-3 

32 2,4’,6 64258-03-3 

33 2’,3,4 

34 2’,3,5 

35 3,3’,4 
PBBs and PBDEs 348 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-1. Chemical Identity of Polybrominated Biphenyl (PBB) and 

Polybrominated Diphenyl Ether (PBDE) Congeners 

CAS No.b IUPAC 

No.a

Compound/ 

substituents Brominated biphenyls (BB)c Brominated diphenyl ethers (BDE)d

36 3,3’,5 

37 3,4,4’ 6683-35-8 

38 3,4,5 115245-08-4 

39 3,4’,5 72416-87-6 

 Tetrabromobiphenyl 40088-45-7 40088-47-9 

40 2,2’,3,3’ 

41 2,2’,3,4 

42 2,2’,3,4’ 

43 2,2’,3,5 

44 2,2’,4,5’ 

45 2,2’,3,6 

46 2,2’,3,6’ 

47 2,2’,4,4' 66115-57-9 

48 2,2’,4,5 

49 2,2’,4,5' 60044-24-8 

50 2,2’,4,6 

51 2,2’,4,6' 97038-95-4 

52 2,2’,5,5' 59080-37-4 

53 2,2’,5,6' 60044-25-9 

54 2,2’,6,6' 97038-96-5 

55 2,3,3’,4 97038-99-8 

56 2,3,3’,4’ 

57 2,3,3’,5 

58 2,3,3’,5’ 

59 2,3,3’,6 

60 2,3,4,4’ 

61 2,3,4,5 115245-09-5 

62 2,3,4,6 115245-10-8 

63 2,3,4’,5 

64 2,3,4’,6 

65 2,3,5,6 

66 2,3’,4,4’ 84303-45-7 

67 2,3’,4,5 

68 2,3’,4,5’ 

69 2,3’,4,6 

70 2,3’,4’,5 59080-38-5 

71 2,3’,4’,6 

72 2,3’,5,5’ 

73 2,3’,5’,6 

74 2,4,4’,5 
PBBs and PBDEs 349 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-1. Chemical Identity of Polybrominated Biphenyl (PBB) and 

Polybrominated Diphenyl Ether (PBDE) Congeners 

CAS No.b IUPAC 

No.a

Compound/ 

substituents Brominated biphenyls (BB)c Brominated diphenyl ethers (BDE)d

75 2,4,4’,6 64258-02-2 

76 2’,3,4,5 

77 3,3’,4,4’ 77102-82-0 

78 3,3’,4,5 

79 3,3’,4,5’ 97038-98-7 

80 3,3’,5,5’ 16400-50-3 

81 3,4,4’,5 59589-92-3 

 Pentabromobiphenyl 56307-79-0 32534-81-9 

82 2,2’,3,3’,4 

83 2,2’,3,3’,5 

84 2,2’,3,3’,6 

85 2,2’,3,4,4’ 

86 2,2’,3,4,5 

87 2,2’,3,4,5’ 

88 2,2’,3,4,6 77910-04-4 

89 2,2’,3,4,6’ 

90 2,2’,3,4’,5 

91 2,2’,3,4’,6 

92 2,2’,3,5,5’ 

93 2,2’,3,5,6 

94 2,2’,3,5,6’ 

95 2,2’,3,5’,6 88700-05-4 

96 2,2’,3,6,6’ 

97 2,2’,3’,4,5 

98 2,2’,3’,4,6 

99 2,2’,4,4’,5 81397-99-1 

100 2,2’,4,4’,6 97038-97-6 

101 2,2’,4,5,5’ 67888-96-4 

102 2,2’,4,5,6’ 80274-92-6 

103 2,2’,4,5’,6 59080-39-6 

104 2,2’,4,6,6’ 97063-75-7 

105 2,3,3’,4,4’ 

106 2,3,3’,4,5 

107 2,3,3’,4’,5 

108 2,3,3’,4,5’ 

109 2,3,3’,4,6 

110 2,3,3’,4’,6 

111 2,3,3’,5,5’ 

112 2,3,3’,5,6 

113 2,3,3’,5’,6 
PBBs and PBDEs 350 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-1. Chemical Identity of Polybrominated Biphenyl (PBB) and 

Polybrominated Diphenyl Ether (PBDE) Congeners 

CAS No.b IUPAC 

No.a

Compound/ 

substituents Brominated biphenyls (BB)c Brominated diphenyl ethers (BDE)d

114 2,3,4,4’,5 96551-70-1 

115 2,3,4,4’,6 

116 2,3,4,5,6 38421-62-4 

117 2,3,4’,5,6 

118 2,3’,4,4’,5 6788-97-5 

119 2,3’,4,4’,6 86029-64-3 

120 2,3’,4,5,5’ 80407-70-1 

121 2,3’,4,5’,6 

122 2’,3,3’,4,5 

123 2’,3,4,4’,5 74114-77-5 

124 2’,3,4,5,5’ 

125 2’,3,4,5,6’ 

126 3,3’,4,4’,5 84303-46-8 

127 3,3’,4,5,5’ 81902-33-2 

 Hexabromobiphenyl 36355-01-8 36483-60-0 

128 2,2’,3,3’,4,4’ 82865-89-2 

129 2,2’3,3’,4,5 

130 2,2’,3,3’,4,5’ 82865-90-5 

131 2,2’,3,3’,4,6 

132 2,2’,3,3’,4,6’ 119264-50-5 

133 2,2’,3,3’,5,5’ 55066-76-7 

134 2,2’,3,3’,5,6 

135 2,2’,3,3’,5,6’ 119264-51-6 

136 2,2’,3,3’,6,6’ 

137 2,2’,3,4,4’,5 81381-52-4 

138 2,2’,3,4,4’,5’ 67888-98-6 

139 2,2’,3,4,4’,6 

140 2,2’,3,4,4’,6’ 

141 2,2’,3,4,5,5’ 120991-47-1 

142 2,2’,3,4,5,6 

143 2,2’,3,4,5,6’ 

144 2,2’,3,4,5’,6 119264-52-7 

145 2,2’,3,4,6,6’ 

146 2,2’,3,4’,5,5’ 

147 2,2’,3,4’,5,6 

148 2,2’,3,4’,5,6’ 

149 2,2’,3,4’,5’,6 69278-59-7 

150 2,2’,3,4’,5,6’ 93261-83-7 

151 2,2’,3,5,5’,6 119264-53-8 

152 2,2’,3,5,6,6’ 
PBBs and PBDEs 351 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-1. Chemical Identity of Polybrominated Biphenyl (PBB) and 

Polybrominated Diphenyl Ether (PBDE) Congeners 

CAS No.b IUPAC 

No.a

Compound/ 

substituents Brominated biphenyls (BB)c Brominated diphenyl ethers (BDE)d

153 2,2’,4,4’,5,5’ 59080-40-9 

154 2,2’,4,4’,5,6’ 36402-15-0 

155 2,2’,4,4’,6,6’ 59261-08-4 

156 2,3,3’,4,4’,5 77607-09-1 

157 2,3,3’,4,4’,5’ 84303-47-9 

158 2,3,3’,4,4’,6 

159 2,3,3’,4,5,5’ 120991-48-2 

160 2,3,3’,4,5,6 

161 2,3,3’,4,5’,6 

162 2,3,3’,4’,5,5’ 

163 2,3,3’,4’,5,6 

164 2,3,3’,4’,5’,6 82865-91-5 

165 2,3,3’,5,5’,6 

166 2,3,4,4’,5,6 

167 2,3’,4,4’,5,5’ 67888-99-7 

168 2,3’,4,4’,5’,6 84303-48-0 

169 3,3’,4,4’,5,5’ 60044-26-0 

 Heptabromobiphenyl 35194-78-6 68928-80-3 

170 2,2’,3,3’,4,4’,5 69278-60-0 

171 2,2’,3,3’,4,4’,6 

172 2,2’,3,3’,4,5,5’ 82865-92-7 

173 2,2’,3,3’,4,5,6 

174 2,2’,3,3’,4,5,6’ 88700-04-3 

175 2,2’,3,3’,4,5’,6 

176 2,2’,3,3’,4,6,6’ 

177 2,2’,3,3’,4’,5,6 

178 2,2’,3,3’,5,5’,6 119264-54-9 

179 2,2’,3,3’,5,6,6’ 

180 2,2’,3,4,4’,5,5’ 67733-52-2 

181 2,2’,3,4,4’,5,6 

182 2,2’,3,4,4’,5,6’ 119264-55-0 

183 2,2’,3,4,4’,5’,6 

184 2,2’,3,4,4’,6,6’ 119264-56-1 

185 2,2’,3,4,5,5’,6 

186 2,2’,3,4,5,6,6’ 119264-57-2 

187 2,2’,3,4’,5,5’,6 84303-49-1 

188 2,2’,3,4’,5,6,6’ 119264-58-3 

189 2,3,3’,4,4’,5,5’ 88700-06-5 

190 2,3,3’,4,4’,5,6 79682-25-0 

191 2,3,3’,4,4’,5’,6 
PBBs and PBDEs 352 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-1. Chemical Identity of Polybrominated Biphenyl (PBB) and 

Polybrominated Diphenyl Ether (PBDE) Congeners 

CAS No.b IUPAC 

No.a

Compound/ 

substituents Brominated biphenyls (BB)c Brominated diphenyl ethers (BDE)d

192 2,3,3’,4,5,5’,6 

193 2,3,3’,4’,5,5’,6 

 Octabromobiphenyl 27858-07-7 32536-52-0 

194 2,2’,3,3’,4,4’,5,5’ 67889-00-3 

195 2,2’,3,3’,4,4’,5,6 

196 2,2’,3,3’,4,4’,5’,6 

197 2,2’,3,3’,4,4’,6,6’ 119264-59-4 

198 2,2’,3,3’,4,5,5’,6 

199 2,2’,3,3’,4,5,6,6’ 

200 2,2’,3,3’,4,5,6,6’ 119264-60-7 

201 2,2’,3,3’,4,5’,6,6’ 69887-11-2 

202 2,2’,3,3’,5,5’,6,6’ 59080-41-0 

203 2,2’,3,4,4’,5,5’,6 

204 2,2’,3,4,4’,5,6,6’ 119264-61-8 

205 2,3,3’,4,4’,5,5’,6 

 Nonabromobiphenyl 27753-52-2 63936-56-1 

206 2,2’,3,3’,4,4’,5,5’,6 69278-62-2 

207 2,2’,3,3’,4,4’,5,6,6’ 119264-62-9 

208 2,2’,3,3’,4,5,5’,6,6’ 119264-63-0 

 Decabromobiphenyl 13654-09-6 1163-19-5 

209 2,2’,3,3’,4,4’,5,5’,6,6’ 13654-09-6 1163-19-5 

a

Ballschmiter and Zell 1980 

b

Not all PBBs have been assigned CAS numbers; with the exception of BDE 209, no CAS numbers were identified 

for the PBDE class. 

c

WHO 1994b 

d

WHO 1994a 
PBBs and PBDEs 353 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-2. Chemical Identity of Selected PBBsa

Characteristic 

Hexabromobiphenyl Octabromobiphenyl Decabromobiphenyl 2,2',4,4'5,5'-Hexabromobiphenyl 

Synonym(s) FireMaster BP-6b; 

FireMaster FF-1b

Bromkal 80b

 Flammex B 10b; 

Adine 0102b

; 

Berkflam B 10b

2,2',4,4'5,5'hexabromo-1,1'-biphenyl 

Registered trade 

name(s) 

FireMaster BP-6; 

FireMaster FF-1 

Bromkal 80 Flammex B 10; 

Adine 0102; 

Berkflam B 10 

None 

Chemical formula C12H4Br6 C12H2Br8 C12Br10 C12H4Br6

Chemical structure 

Identification 

numbers: 

 

CAS registry 59536-65-1 (BP-6); 

67774-32-7 (FF-1); 

36355-01-8 (hexabromo mixture) 

27858-07-7 (octobromo mixture) 

61288-13-9 

(Bromkal 80) 

13654-09-6 (pure 

and technical) 

59080-40-9 

NIOSH RTECS LK 5060000 (BP-6); 

LK 5065000 (FF-1) 

DV 570000 (octabromo mixture) No data No data 

EPA hazardous 

waste 

No data No data No data No data 

OHM/TADS No data No data No data No data 

DOT/UN/NA/IMCO 

shipping 

No data No data No data No data 

HSDB No data No data No data 2913 

NCI No data No data No data No data 

a

All information obtained from IARC 1986 except where noted. 

b

These are mixtures of compounds, and their compositions are given in the text. 

CAS = Chemical Abstracts Services; DOT/UN/NA/IMCO = Department of Transportation/United Nations/North 

America/International Maritime Dangerous Goods Code; EPA = Environmental Protection Agency; 

HSDB = Hazardous Substances Data Bank; NCI = National Cancer Institute; NIOSH = National Institute for 

Occupational Safety and Health; OHM/TADS = Oil and Hazardous Materials/Technical Assistance Data System; 

RTECS = Registry of Toxic Effects of Chemical Substances
PBBs and PBDEs 354 

6. CHEMICAL AND PHYSICAL INFORMATION 

It can be seen from the structure that a large number of brominated compounds are possible. The 

209 possible compounds for PBDEs are called “congeners”. However, the number of PBDE congeners 

that actually exist in commercial PBDE mixtures are much less compared to PCBs. Typically, only a 

subset of the 209 possible congeners is observed for PBDEs. PBDEs can also be categorized by degree of 

bromination. The term “homolog” is used to refer to all PBDEs with the same number of bromines (e.g., 

tribromodiphenyl ether refers to PBDEs containing only three bromine atoms). Based on the number of 

bromine substituents, there are 10 homologous groups of PBDEs (monobrominated through 

decabrominated). Each homologous group contains one or more congeners. The mono-, di-, tri-, tetra-, 

penta-, hexa-, hepta-, octa-, nona-, and decabromo-congeners can exist in 3, 12, 24, 42, 46, 42, 24, 12, 3, 

and 1 forms, respectively. Homologs with different substitution patterns are referred to as isomers. For 

example, the group of dibromodiphenyl ether homologs contains 12 isomers. The numbering system for 

PBDEs is also shown above. The structures of representative PBB and PBDE molecules appear similar 

when drawn in one dimension. However, there are important three-dimensional differences in their 

structures due to the ether linkage and location/number of halogen atoms. The ortho positions of the 

aromatic rings must be nonhalogen-substituted for a diphenyl ether molecule to assume a planar or near 

planar configuration. Halogen substitution of the diphenyl ether molecule in the ortho position (2,2’,6,6’) 

will force the aromatic rings orthogonal to one another (e.g., the phenyl rings will be positioned in space 

with a dihedral angle >0°). This is particularly evident for decabromodiphenyl ether, which is predicted 

to have a dihedral angle of ca. 90° and a high barrier to rotation around the ether linkage preventing this 

molecule from assuming a planar configuration. The benzene rings of non-ortho substituted PBDEs may 

assume a small dihedral angle (in which the dihedral angle is small, but >0°) or “near” planar 

configuration. These molecules are referred to as planar or coplanar congeners (Hardy 2002). 

Like PCBs, the 209 congeners for PBDEs are arranged in ascending numerical order using a numbering 

system developed by Ballschmiter and Zell (1980) that follow the IUPAC rules of substituent 

characterization in biphenyls. The resulting numbers assigned by Ballschmiter and Zell (which are also 

referred to as congener, IUPAC, or BZ numbers) are widely used for identifying individual congeners of 

PBDEs. For example, the PBDE congener, 2,2',4,4'-tetrabromodiphenyl ether may be referred to as BDE 

47 in this document. The identities of several PBDE congeners are shown in Table 6-1 (WHO 1994a, 

1994b). 

In the United States, Albemarle Corporation and Great Lakes Chemical Corporation market mixtures of 

PBDEs under trade names (e.g., DE-60F, DE-61, DE-62, and DE-71 for pentaBDE mixtures; DE-79 for 

octaBDE mixtures; and DE 83R, Saytex 102E for decaBDE mixtures). There are also several trade 
PBBs and PBDEs 355 

6. CHEMICAL AND PHYSICAL INFORMATION 

names used by producers from Europe and Japan for the BDE mixtures. The chemical identities of 

commercial mixtures of penta-, octa-, and decabromodiphenyl ethers are listed in Table 6-3 (WHO 

1994a). 

Various synonyms and abbreviations for PBDEs exist in the literature and are shown below: 

polybrominated biphenyl ethers = Polybromobiphenyl ethers = PBBE 

polybrominated biphenyl oxides = Polybromobiphenyl oxides = PBBEs 

polybrominated diphenyl ethers = Polybromodiphenyl ethers = PBDEs or PBDPEs 

polybrominated diphenyl oxides = Polybromodiphenyl oxides = PBDOs or PBDPOs 

For consistency in this document, polybrominated diphenyl ethers or PBDEs will be used to identify this 

class of chemicals. The PBDE homologs are abbreviated as follows in this document: 

dibromodiphenyl ether = DiBDE = diBDE 

tribromodiphenyl ether = TrBDE = triBDE 

tetrabromodiphenyl ether = TeBDE = tetraBDE 

pentabromodiphenyl ether = PeBDE = pentaBDE 

hexabromodiphenyl ether = HxBDE = hexaBDE 

heptabromodiphenyl ether = HpBDE = heptaBDE 

octabromodiphenyl ether = OBDE = octaBDE 

nonabromodiphenyl ether = NoBDE = nonaBDE 

decabromodiphenyl ether = DeBDE = decaBDE 

6.2 PHYSICAL AND CHEMICAL PROPERTIES 

Polybrominated Biphenyls. Information found in the literature regarding the physical and chemical 

properties of hexabromobiphenyl, octabromobiphenyl, decabromobiphenyl, and BB 153 is presented in 

Table 6-4. The data for the properties listed in Table 6-4 may not be reliable because products of 

questionable purity were used by earlier investigators to derive them. For example, the water solubility of 

hexabromobiphenyl (Neufeld et al. 1977) was reported to be the same as that of FireMaster FF-1 (Getty et 

al. 1977), although FireMaster FF-1 contained only 84.4% (Robertson et al. 1983b) hexabrominated 

biphenyls. However, recent physical and chemical property data have been reported for 

hexabromobiphenyl in Tittlemier et al. (2002). 

Of the 209 possible congeners of PBBs, only about 42 have been synthesized in pure form even on a 

laboratory scale (Sundstrom et al. 1976b). The PBBs produced for commercial use were mixtures of 

PBBs with other non-PBB impurities. The technical products were FireMaster BP-6, FireMaster FF-1, 

Bromkal 80, and Flammex B 10 (or Adine 0102 or Berkflam B 10) (IARC 1986). FireMaster FF-1, a 
PBBs and PBDEs 356 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-3. Chemical Identity of Technical PBDEs 

Characteristic

Pentabromodiphenyl 

ether

Octabromodiphenyl 

ether Decabromodiphenyl ether

Synonym(s) Pentabromodiphenyl 

ether; pentabromodiphenyl oxide; pentabromobiphenyl oxide; 

benzene, 1,1-oxybis, 

pentabromo derivative 

Octabromodiphenyl 

ether; Octabromodiphenyl oxide; octabromobiphenyl oxide; 

benzene, octabromo 

derivative; phenyl ether, 

octabromo derivative 

Decabromodiphenyl ether; decabromodiphenyl oxide; decabromobiphenyl oxide; benzene, 1,1’-oxybis(2,3,5,6,-penta-bromo-) ether, bis-

(pentabromophenyl); 

Registered trade 

name 

DE 71; Bromkal 70-5 

DE; FR 1205/1215; 

Bromkal 70; Bromkal 

G1; Pentabromprop; 

Tardex 50; Tardex 50 L; 

Saytex 115 

Bromkal 7908DE; DE 

79; FR 143; Tardex 80; 

FR 1208; Adine 404; 

Saytex 111 

FR-300 BA; DE-83-RTM; Saytex 

102; Saytex 102E; FR-1210; Adine 

505; AFR 1021; Berkflam B10E; 

BR55N; Bromkal 81; Bromkal 

82-ODE; Bromkal 83-10 DE; Caliban 

F/R-P 39P; Caliban F/R-P 44; 

Chemflam 011; DE 83; DP 10F; EB 

10FP; EBR 700; Flame Cut BR 100; 

FR P-39; BR 100; FR 330BA; FR 

P-39; FRP 53; FR-PE; FR-PE(H); 

Planelon DB 100; Tardex 100; 

NC-1085; HFO-102; Hexcel PF1; 

Phoscon Br-250 

Chemical 

formula 

C12H10-yBryO 

where y=4–6 

C12H10-yBryO 

where y=6–9 

C12Br10O 

Chemical 

structure 

 

Identification numbers: 

 CAS registry 32534-81-9 32536-52-0 1163-19-5 

 NIOSH RTECS No data No data No data 

 EPA hazardous 

waste 

No data No data No data 

 OHM/TADS No data No data No data 

 DOT/UN/IMCO 

shipping 

No data No data No data 

 HSDB 7109 7110 2911 

 NCI No data No data No data 

Source: WHO 1994a 

CAS = Chemical Abstracts Services; DOT/UN/NA/IMCO = Department of Transportation/United Nations/North 

America/International Maritime Dangerous Goods Code; EPA = Environmental Protection Agency; 

HSDB = Hazardous Substances Data Bank; NCI = National Cancer Institute; NIOSH = National Institute for 

Occupational Safety and Health; OHM/TADS = Oil and Hazardous Materials/Technical Assistance Data System; 

PBB = polybrominated biphenyl; RTECS = Registry of Toxic Effects of Chemical Substances
PBBs and PBDEs 357 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-4. Physical and Chemical Properties of Selected PBBsa

Property Hexabromobiphenyl

Octabromobiphenyl Decabromobiphenyl 2,2',4,4',5,5'-Hexabromobiphenyl 

Molecular weight 627.4 785.2 943.1 627.4 

Color White White White White 

Physical state Solid Solid Solid Solid 

Melting point 72 °C 200–250 °C; 367–

367 °Cb

 (for 

industrial product)

380–386 °C No data 

Boiling point No data No data No data No data 

Density No data No data No data No data 

Odor No data No data No data No data 

Odor threshold: 

 Water No data No data No data No data 

 Air No data No data No data No data 

Solubility: 

 Water 11 μg/L; 3 μg/Lc 20–30 μg/L Insoluble 11 μg/Ld

 Organic 

solvent(s) 

Soluble in acetone, 

benzene 

Soluble in 

methylene 

chloride, benzene 

Moderately soluble 

in chlorobenzene, 

o-xylene 

Acetone (6 weight 

percent); benzene 

(75 weight percent)c

Partition coefficients: 

 Log Kow 6.39e

 5.53 8.58f 9.10 (estimated)d

 Log Koc 3.33–3.87g

 No data No data 5.088d

Vapor pressure 5.2x10-8 mmHg at 

25 °Ch

; 5.6x10-6 mm 

Hg (liquid subcooled)c 

7x10-11 mmHg at 

28 °Ci

No data 7.6x10-5 mm Hg at 

90 °Cd

Henry's law 

constant 

3.9x10-6 atm-m3

/molj; 

1.38x10-6 atm-m3

/molc

No data No data 5.7x10-3 atm-m3/mold

Autoignition 

temperature 

No data No data No data No data 

Flashpoint No data No data No data No data 

Flammability limits No data No data No data No data 

Conversion 

factors 

Since these compounds exist in the particle phase in the ambient atmosphere, the 

concentrations in air are expressed in weight per unit volume of the air. 

Explosive limits No data No data No data No data 

a

All information obtained from IARC (1978) and Norris et al. (1973) unless otherwise noted. 

b

Sundstrom et al. 1976 

c

Tittlemier et al. 2002 

d

Hardy (2002) 

e

Doucette and Andren 1988 

f

The values for 2,2'4,4'6,6'- and 2,2',3,3',4,4'-hexabromobiphenyl are given as 7.20 (Chessells et al. 1992) and 

8.09 (Anliker et al. 1988), respectively. 

g

Estimated from the Freundlich adsorption constants given by Jacobs et al. (1978). h

Jacobs et al. 1976 

i

Waritz et al. 1977 

j

Estimated from the ratio of vapor pressure and water solubility. 
PBBs and PBDEs 358 

6. CHEMICAL AND PHYSICAL INFORMATION 

white powder, was made by grinding brown flakes of FireMaster BP-6 and adding 2% calcium silicate as 

an anticaking agent (Fries 1985b). The exact composition of FireMaster BP-6 or FireMaster FF-1 seems 

to have varied between and within batches (Sundstrom et al. 1976a). Table 6-5 provides the 

concentrations of the PBB congeners in FireMaster FF-1 and FireMaster BP-6. 

An interesting feature of commercial FireMaster FF-1 and FireMaster BP-6 is that they contain >50% of 

the congener BB 153. The second most abundant congener is 2,2',3,4,4',5,5'-heptabromobiphenyl 

(BB 180). A detailed analysis of FireMaster BP-6 (lot 7062) was able to separate 22 congeners of PBBs 

that included four tri, five tetra, three penta, seven hexa, and three hepta congeners of PBBs (Robertson et 

al. 1983b, 1984b). The coplanar and toxic congeners 3,3',4,4'-tetra-, 3,3',4,4',5-penta-, and 3,3',4,4',5,5'-

hexabrominated biphenyls were found at abundances of 0.159, 0.079, and 0.294%, respectively (Orti et 

al. 1983; Robertson et al. 1983b). In addition to the 22 congeners, other investigators have identified 

2,2',3,3',4,4',5,6'-octa-, 2,2',3,3',4,4',5,5'-octa-, 2,2',3,3',4,4',5,5',6-nona-, and decabromobiphenyl in 

commercial PBBs (Moore et al. 1978). Other impurities detected in FireMaster FF-1 and FireMaster 

BP-6 were tetra-, penta-, and hexabromonaphthalene (Di Carlo et al. 1978); however, at a detection limit 

of 0.5 ppm, brominated dioxins and dibenzofurans were not detected in commercial FireMaster FF-1 or 

FireMaster BP-6 (Hass et al. 1978). 

Commercial octabromobiphenyl (Bromkal 80) contained at least four compounds. Assays of two 

commercial octabromobiphenyls showed the following compositions: 1.0–1.8% heptabromobiphenyl, 

33.0–45.2% octabromobiphenyl, 47.4–60.0% nonabromobiphenyl, and 5.7–6.0% decabromobiphenyl 

(Norris et al. 1973; Waritz et al. 1977). Notably, the major component of commercial octabromobiphenyl 

was nonabromobiphenyl, and not octabromobiphenyl. Commercial decabromobiphenyl (Flammex B 10) 

contained 96.8% decabromobiphenyl, 2.9% nonabromobiphenyl, and 0.3% octabromobiphenyl (Di Carlo 

et al. 1978). 

Pyrolysis of FireMaster BP-6 in the temperature range of 600–900 °C in the absence of oxygen produced 

bromobenzenes and brominated biphenyls as key products, but no brominated dioxins and dibenzofurans 

(Thoma and Hutzinger 1987; Thoma et al. 1987). Thermolysis of FireMaster BP-6 between 400 and 

600 °C in the presence of air produced 2,3,7,8-tetrabromodibenzofuran in the percent (1%=10 g/kg) range 

(Rappe and Buser 1980). Pyrolysis of FireMaster BP-6 in an open quartz tube at 800 °C produced 0.48–

1.49 g/kg 2,3,7,8-TCDD equivalent levels of polybrominated dibenzofurans (Zacharewski et al. 1988). 

FireMaster BP-6 hydrolyzed when refluxed with 2% potassium hydroxide in ethanol, but the possible rate 
PBBs and PBDEs 359 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-5. Identified PBB Congeners in FireMaster® BP-6 and FireMaster® FF-1 

Percent composition of 

IUPAC No.a Structure FireMaster BP-6 FireMaster FF-1 References 

Dibromobiphenyls 

4 2,2’- 0.020 Moore et al. 1979 

Tribromobiphenyls 

18 2,2’,5- 0.050 Robertson et al. 1984 

26 2,3’,5- 0.024 

31 2,4’,5- 0.015 

37 3,4,4’- 0.021 

Tetrabromobiphenyls 

49 2,2’,4,5’- 0.025 

52 2,2’,5,5’- 0.052 

66 2,3’,4,4’- 0.028 

70 2,3’,4’,5- 0.017 

77 3,3’,4,4’- <0.080 Orti et al. 1983 

 0.159 Robertson et al. 1984 

Pentabromobiphenyls 

95 2,2’,3,5’,6- 0.020 Orti et al. 1983 

99 2,2’,4,4’,5- <0.08 

101 2,2’,4,5,5’- 2.69 Robertson et al. 1984 

 4.50 3.70 Aust et al. 1981 

 1.54 Orti et al. 1983 

 2.60 Krüger 1988 

118 2,3’,4,4’,5- 2.94 Robertson et al. 1984 

 0.70 Robertson et al. 1984 

 3.20 Krüger 1988 

126 3,3’,4,4’,5- <0.01 

 0.079 Robertson et al. 1984 

Hexabromobiphenyls 

132 2,2’3,3’,4,6’- 1 Krüger 1988 

138 2,2’,3,4,4’,5’- 12.3 Robertson et al. 1984 

 12 8.6 Aust et al. 1981 

 5.23 Orti et al. 1983 

 10.6 Krüger 1988 

149 2,2’,3,4’,5’,6- 2.24 Robertson et al. 1984 

 1.40 1.30 Aust et al. 1981 

 0.78 Orti et al. 1983 

153 2,2’,4,4’,5,5’- 53.9 Robertson et al. 1984 

 47.8 47.1 Aust et al. 1981 

 55.2 Orti et al. 1983 

 58.5 Krüger 1988 
PBBs and PBDEs 360 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-5. Identified PBB Congeners in FireMaster® BP-6 and FireMaster® FF-1 

Percent composition of 

IUPAC No.a Structure FireMaster BP-6 FireMaster FF-1 References 

155 2,2’4,4’,6,6’- 0.5 

156 2,3,3’,4,4’,5- 0.980 Robertson et al. 1984 

 5.0 Aust et al. 1981 

 0.37 Orti et al. 1983 

 1.0 Krüger 1988 

157 2,3,3’,4,4’,5’- 0.05 Orti et al. 1983 

 0.526 Robertson et al. 1984 

 0.5` Krüger 1988 

167 2,3’,4,4’,5,5’- 5.5 3.3 Aust et al. 1981 

 3.37 Orti et al. 1983 

 <0.3 

 7.95 Robertson et al. 1984 

 5.5 Krüger 1988 

169 3,3’,4,4’,5,5’- 0.294 Robertson et al. 1984 

Heptabromobiphenyls 

170 2,2’,3,3’,4,4’,5- 0.256 

 1.1 1.5 Aust et al. 1981 

 1.66 Orti et al. 1983 

 2.4 Krüger 1988 

172 2,2’,3,3’,4,5,5’- <0.30 Orti et al. 1983 

174 2,2’,3,3’,4,5,6’- 0.24 

178 2,2’,3,3’,5,5’,6- 0.3 Krüger 1988 

180 2,2’,3,4,4’,5,5’ 6.97 Robertson et al. 1984 

 24.7 Aust et al. 1981 

 23.5 Orti et al. 1983 

187 2,2’,3,4’,5,5’,6- 0.392 Robertson et al. 1984 

 1.0 Krüger 1988 

189 2,3,3’,4,4’,5,5’- 0.51 Orti et al. 1983 

Octabromobiphenyls 

194 2,2’,3,3’,4,4’,5,5’- 0.9 2.4 Aust et al. 1981 

 1.65 Orti et al. 1983 

196 2,2’,3,3’,4,4’,5,6’- Moore et al. 1980 

201 2,2’,3,3’,4,5,5’,6’- Orti et al. 1983 

203 2,2’,3,4,4’,5,5’,6- 

Source: WHO 1994b 

a

Ballschmiter and Zell 1980 
PBBs and PBDEs 361 

6. CHEMICAL AND PHYSICAL INFORMATION 

of PBB hydrolysis under much milder environmental conditions remains unknown (Pomerantz et al. 

1978). 

Hexabromonapthalene has been identified as a toxic contaminant of Firemaster BP-6 or FF-1 at 

concentration levels of approximately 150 ppm (Birnbaum et al. 1983). Previously reported to be a single 

compound, hexabromonapthalene was shown to be a 60:40 mixture of 1,2,3,4,6,7-hexabromonaphthalene 

and 2,3,4,5,6,7-hexabromonaphthalene. 

Polybrominated Diphenyl Ethers. Information found in the literature regarding the physical and 

chemical properties of selected technical PBDE mixtures is presented in Table 6-6. Recent information 

regarding the vapor pressure, water solubility, Henry’s Law constant, and log Kow of some PBDE 

congeners is presented in Table 6-7. 

Commercially available product mixtures of PBDEs (see Table 6-3) are not pure substances, but instead 

are mixtures of congeners. For example, the commercial mixture pentabromodiphenyl ether denotes the 

main component of the mixture contains the pentabromodiphenyl ether homolog. However, the 

commercial pentaBDE mixture actually contains tetrabromodiphenyl ether (24–38%) and 

pentabromodiphenyl ether (50–62%) homologs with small amounts of hexabromodiphenyl ether (4–8%) 

and trace amounts of tribromodiphenyl ether (0–1%) homologs. In this document, the commercial 

mixture of pentaBDE may be called “the commercial pentaBDE mixture,” “technical pentaBDE,” or 

“technical PeBDE” to distinguish this mixture of homologs from the pentaBDE homolog which refers to 

polybrominated diphenyl ethers with only five bromine atoms (see Section 6.1). Commercial 

octabromodiphenyl ether is a mixture of hexa-, hepta-, octa-, and nonabrominated diphenyl ether 

homologs with trace amounts of decabromodiphenyl ether (i.e., BDE 209). In this document, the 

commercial mixture of octaBDE may be called “the commercial octaBDE mixture,” “technical 

octaBDE,” or “technical OBDE” to distinguish this mixture of different homologs from the octaBDE 

homolog, which refers to polybrominated diphenyl ethers with only eight bromine atoms (see 

Section 6.1). The composition of commercial decabromodiphenyl ether is 97% of the 

decabromodiphenyl ether (i.e., BDE 209); the remainder is nonabromodiphenyl ether homologs and trace 

amounts of octabromodiphenyl ether homologs (WHO 1994a). In this document, commercial 

decabromodiphenyl ether may be called “the commercial decaBDE mixture,” “technical decaBDE,” or 

“technical DeBDE” which represents 97% BDE 209 congener with 3% nona- and octaBDE homolog 

impurities. The composition of commercial octaBDE (e.g., DE-79) is <10% of nonaBDE, <33% of 

octaBDE, <45% of heptaBDE, and <12% of hexaBDE (i.e., 2% of BDE 154 and 14% of BDE 153) 
PBBs and PBDEs 362 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-6. Physical and Chemical Properties of Technical PBDE Mixtures 

Property

Pentabromodiphenyl 

ether

Octabromodiphenyl 

ether Decabromodiphenyl ether

Molecular weight Mixture Mixture 959.22a

Color Clear, amber to pale 

yellowa

Off-whitea

 Off-whitea

Physical state Highly viscous liquid Powder Powdera

Melting point -7 to -3 °C (commercial)b

 85–89 °C (commercial)c; 

200 °C (range, 167–

257)a

; 79–87 °Ca; 170–

220 °Ca

290–306 °Ca

Boiling point >300 °C (decomposition 

starts above 200 °C)a,b

Decomposes at >330 °C 

(commercial)c 

Decomposes at >320, >400, 

and 425 °Ca

Density (g/mL) 2.28 at 25 °Ca

; 2.25–2.28b 2.76a; 2.8 (commercial)c 3.0a; 3.25a

Odor No data Fainta

 Odorlessa

Odor threshold: 

 Water No data No data Not applicable 

 Air No data No data Not applicable 

Solubility: 

 Water 13.3 μg/L (commercial)b,d; 

2.4 μg/L (pentabromodiphenyl ether component)b

; 

10.9 μg/L (tetrabromodiphenyl ether component)b 

<1 ppb at 25 °C (commercial)c

; 1.98 μg/L 

(heptabromodiphenyl 

ether component)c

<0.1 μg/Lg 

 Organic solvent(s) 10 g/kg methanol; miscible 

in toluened

acetone (20 g/L); 

benzene (200 g/L); 

methanol (2 g/L) all at 

25 °Ca

acetone (0.05%), benzene 

(0.48%), methylene bromide 

(0.42%), xylene (0.87%), 

and toluene (0.2%)e 

Partition coefficients: 

 Log Kow 6.64–6.97d

; 6.57 

(commercial)b 

6.29 (commercial)c 6.265e 

 Log Koc 4.89-5.10f 5.92-6.22f 6.80f

Vapor pressure 2.2x10-7–5.5x10-7 mm Hg 

at 25 °Cd

; 3.5x10-7 mm Hg 

(commercial)b 

9.0x10-10–1.7x10-9 mm 

Hg at 25 °Cd

; 4.9x10-8

mm Hg at 21 °C 

(commercial)c 

3.2x10-8 mm Hgg; 

3.47x10-8 mm Hge 

Henry’s Law constant 

(atm-m3

/mole) 

1.2x10-5h; 1.2x10-6f; 

3.5x10-6f

7.5x10-8f; 2.6x10-7f 1.62x10-6h; 1.93x10-8d; 

1.2x10-8f; 4.4x10-8f

Autoignition 

temperature 

Decomposes above 

200 °Cb,d

Decomposes above 

330 °C (commercial)c 

Not applicablea

Flashpoint No data No data None 

Flammability limits Not applicable (flame 

retardant)b,d

Not applicable (flame 

retardant)c 

Non-flammablea
PBBs and PBDEs 363 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-6. Physical and Chemical Properties of Technical PBDE Mixtures 

Property

Pentabromodiphenyl 

ether

Octabromodiphenyl 

ether Decabromodiphenyl ether

Conversion factors 1 ppm=23.48 mg/m3 at 

20 °Cd

No data No data 

Explosive limits Noneb,g Nonec No data 

a WHO 1994a 

b

ENVIRON 2003a 

c

ENVIRON 2003b 

d

EU 2001 

e

American Chemistry Council 2002 

f

Estimated values were calculated using EPIWIN v3.10 (EPA 2001). 

g

Hardy 2002 

h

Estimate value was calculated using vapor pressure and water solubility values in table. 
PBBs and PBDEs 364 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-7. Physical and Chemical Properties of Some PBDE Congenersa,b

Congener 

Vapor pressure 

(mm Hg)a,c 

Water solubility 

(μg/L)a 

Henry’s Law constant 

(atm m3/mol)a Log Kowb 

BDE-3 1.94x10-3 – – – 

BDE-15 1.30x10-4 130 2.07254x10-4 – 

BDE-17 – – – 5.74 

BDE-28 1.64x10-5 70 5.03331x10-5 5.94 

BDE-47 1.40x10-6 15 1.48038x10-5 6.81 

BDE-66 9.15x10-7 18 4.93461x10-6 – 

BDE-77 5.09x10-7 6 1.18431x10-5 – 

BDE-85 7.40x10-8 6 1.08562x10-6 – 

BDE-99 1.32x10-7 9 2.26992x10-6 7.32 

BDE-100 2.15x10-7 40 6.80977x10-7 7.24 

BDE-138 1.19x10-8 – – – 

BDE-153 1.57x10-8 1 6.61238x10-7 7.90 

BDE-154 2.85x10-8 1 2.36862x10-6 7.82 

BDE-183 3.51x10-9 2 7.30323x10-8 8.27 

BDE-190 2.12x10-9 – – – 

a

Tittlemier et al. 2002 

b

Braekvelt et al. 2003 

c

liquid sub-cooled vapor pressures 

– = No data reported; BDE = brominated diphenyl ether 
PBBs and PBDEs 365 

6. CHEMICAL AND PHYSICAL INFORMATION 

(ENVIRON 2003b). The composition of commercial pentaBDE (e.g., DE-71) is 4–12% of hexaBDE 

(i.e., 4% BDE 154 and 6% BDE 153), 50–62% of pentaBDE (i.e., 8% of BDE 100, 43% of BDE 99), 24–

38% of tetraBDE (i.e., <1% of BDE 66, <1% of BDE 49, 28% of BDE 47), and 0–1% of triBDE (i.e., 

>1% of BDE 28/33) (ENVIRON 2003b). The compositions of commercial product mixtures of PBDEs 

(e.g., technical penta-, octa-, and decaBDE) are given in Table 6-8. The chemical composition of 

commercial pentaBDE product has been shown to vary over the past 10–20 years and between different 

commercial products (ENVIRON 2003a). 

Trace analysis of these commercial mixtures for 15-different 2,3,7,8-substituted brominate dibenzop-dioxins and dibenzofurans revealed no detectable amounts of these substances (Hardy 2002). The 

commercial decaBDE product has been analyzed for trace quantities of 15 2,3,7,8-substituted 

polybrominated-p-dibenzodioxins (PBDDs) and polybrominated dibenzofurans (PBDFs). None of the 

analytes were present at or above the quantization limits established under an EPA test rule (BFRIP 

2002). While in today’s commercial PBDE samples, there are not measurable quantities of 

PBDDs/PBDFs, there are some materials that have reported quantifiable levels of these contaminants. 

For example, hexabromodibenzofurans have been detected in commercial decaBDE mixtures at 

concentrations as high as 200 μg/kg. In other PBDE mixtures (e.g., tetra- to hexaBDEs), the sum of 

tetra-, penta-, and hexabromodibenzofurans were reported at a concentrations of 8,000 μg/kg. In addition, 

tetra- and pentabromo- p-dibenzodioxins have been measured in commercial decaBDE at concentrations 

of 0.05 and 0.35 μg/kg, respectively (WHO 1998). 

When pyrolyzed up to 900 °C, PBDEs may produce PBDFs and PBDDs (Buser 1986; EU 2001). The 

amount of PCDFs and PBDDs formed depends upon the conditions of pyrolysis. For example, 

2,3,7,8-tetrabromodibenzofuran in ppm concentrations can be generated during pyrolysis of 

decabromodiphenyl ether (decaBDE) in the temperature range of 400–700 °C (Bieniek et al. 1989). 

PBDFs may also be produced during the pyrolysis of polymers containing PBDEs as flame retardants 

(Brenner and Knies 1993; Dumler et al. 1989a, 1990; Lenoir et al. 1994). However, studies performed in 

the late 1980’s may have suffered from analytical methods that could not differentiate between PBDD/Fs 

formed (e.g., 2,3,7,8-substituted congeners) and decaBDE which might have artificially elevate levels of 

PBDFs detected (Ranken et al. 1994). 
PBBs and PBDEs 366 

6. CHEMICAL AND PHYSICAL INFORMATION 

Table 6-8. Composition of Commercial Brominated Diphenyl Ethers 

Great Lakes 

Commercial pentaBDE 

product 

Commercial octaBDE 

product 

Commercial 

decaBDE product 

2002 

Late 1970s –

early 1980s 2002 

Late 1970s – 

early 1980s 2002 

BDE 

homologue 

group 

BDE 

congener

a DE-71b Product DE-79b Product DE-83R and DE-83b

DecaBDE 

(1 congener) 

BDE-209 – 0.8 <0.70% 1.6% >98% 

NonaBDE 

(3 congeners) 

– – 0.2 <10% 13.0% – 

OctaBDE 

(12 congeners) 

– – 0.3 <33% 30.7% – 

HeptaBDE 

(24 congeners) 

– – 2.6 <45% 45.1% – 

 4–12% 13.3 <12% 8.5% – 

BDE-154 (4%) (2%) 

HexaBDE 

(42 congeners) 

BDE-153 (6%) (14%) 

 50–62% 58.1 <0.50% 1.1% – 

BDE-100 (8%) (ND) 

BDE-99 (43%) (<1%) 

PentaBDE 

(46 congeners) 

BDE-85 (–) (–) 

 24–38% 24.6 – – – 

BDE-66 (<1%) (ND) 

BDE-49 (<1%) (ND) 

TetraBDE 

(42 congeners) 

BDE-47 (28%) (<1%) 

TriBDE 0–1% – – – – 

(24 congeners) BDE-28/33 (>1%) (ND) 

Source: ENVIRON 2003b 

aIUPAC numbering system 

bAs currently produced 

– = Data not reported; ( ) = % concentration of congener; BDE = bromodiphenyl ether; ND = Not 

detected 
PBBs and PBDEs 367 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

7.1 PRODUCTION 

Polybrominated Biphenyls. The commercial production of polybrominated biphenyls (PBBs) generally 

involves bromination of biphenyl, a process involving a much more specific reaction and that produces a 

smaller number of product mixtures than chlorination (Sundstrom et al. 1976a). In one process, biphenyl 

is brominated with 0–20% stoichiometric excess of bromine chloride (e.g., slightly more than 10 mol of 

bromine chloride may be reacted with 1 mol of biphenyl to obtain decabromobiphenyl) in the presence of 

iron or a Friedel-Crafts catalyst (e.g., aluminum chloride). In another process, biphenyl is dissolved in 

ethylene bromide solvent and reacted with bromine in the presence of a catalyst (either aluminum 

chloride or bromide) (Neufeld et al. 1977). Research quantities of PBBs can be synthesized by the diazo 

coupling of brominated aniline with an excess of the corresponding bromobenzene. For example, 

2,3,3',4,4',5'-hexabromobiphenyl can be synthesized by the diazo coupling of 3,4,5-tribromoaniline with 

1,2,3-tribromobenzene (Kubiczak et al. 1989; Robertson et al. 1983b). Methods for laboratory scale 

synthesis of 42 congeners of brominated biphenyls are also available (Sundstrom et al. 1976b). 

The commercial production of PBBs began in 1970. Approximately 13.3 million pounds of PBBs were 

produced in the United States from 1970 to 1976. Only three commercial PBB products were 

manufactured (i.e., hexabromobiphenyl, octabromobiphenyl, and decabromobiphenyl) and these three 

products were based on a limited number of congeners (Hardy 2002b). Hexabromobiphenyl constituted 

about 11.8 million pounds (ca 88%) and octa- and decabromobiphenyl constituted ≈1.5 million pounds 

together of this total (Neufeld et al. 1977). Over 98% of the hexabromobiphenyl was produced as 

FireMaster BP-6 and the residual as FireMaster FF-1 (Hesse and Powers 1978). Michigan Chemical 

Corporation, St. Louis, Michigan, the sole producer of hexabromobiphenyl in the United States, stopped 

producing this PBB in 1975. White Chemical Co., Bayonne, New Jersey, and Hexcel Corporation, 

Sayreville, New Jersey, manufactured octa- and decabromobiphenyl in the United States until 1979 

(IARC 1986; Neufeld et al. 1977). Shortly after the 1973–1974 agriculture contamination episode in 

Michigan (see Section 5.2), PBB production in the United States was voluntarily discontinued (Hardy 

2000); PBBs are no longer produced in the United States (SRI 2001). Re-initiation of manufacture of 

PBBs requires approval from the EPA. Production of decaPBB in Great Britain was discontinued in 1977 

and highly brominated PBBs were produced in Germany until mid-1985. Until the year 2000, the only 
PBBs and PBDEs 368 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

PBB in commercial production was decabromobiphenyl, which was manufactured by one company 

(Atochem) in France (Hardy 2000). 

Polybrominated Diphenyl Ethers. The commercial production of PBDEs generally involves bromination 

of diphenyl oxide to varying degrees. The degree of bromination is controlled either through stiochiometry or through control of reaction kinetics (Pettigrew 1993). Technical decabromodiphenyl ether is 

manufactured by bromination of diphenyl oxide in the presence of a Friedel-Crafts catalyst (e.g., 

aluminum bromide) and excess bromine. Technical decabromodiphenyl ether may also be produced at 

atmospheric pressure by dissolving diphenyl oxide and bromine in ethylene dibromide in the presence of 

a Friedel-Crafts catalyst (e.g., aluminum bromide). The use of bromine in an organic solvent requires 

long reaction times, results in low productivity per volume, and necessitates recycling of the solvent. 

These limitations have led to the use of bromine as both the reactant and the solvent for this process 

(Dagani and Sanders 1985). 

The commercial production of PBDEs began in late 1970s (WHO 1994a). As of 2003, in the United 

States, penta- and octabromodiphenyl ethers (pentaBDE and octaBDE) are produced by the Great Lakes 

Chemical Corporation in El Dorado, Arkansas (SRI 2002). According to SRI (2002), as of 2003, 

technical decabromodiphenyl ether (decaBDE) is produced by Albemarle Corporation in Magnolia, 

Arkansas. Technical PentaBDE is produced commercially in the United States as Great Lakes DE-60F, 

DE-61, DE-62, and DE-71. Technical octaBDE is produced commercially in the United States as DE-79. 

Technical decaBDE (BDE 209) is produced commercially in the United States as DE-83R and Saytex 

102E. Dead Sea Bromines/Eurobrome and Tosoh are currently producing PBDEs in the Netherlands and 

Japan, respectively (WHO 1994a). No current estimates of PBDE production quantities were located 

(SRI 2002). 

In 2001, the total market demand for PBDEs in the Americas was 33,100 metric tons (see Table 7-1). 

Technical decaBDE constituted about 24,500 metric tons (74%), while technical mixtures of octaBDEs 

and pentaBDEs were 1,500 and 7,100 metric tons (4 and 22 %) of this total, respectively (BSEF 2003). 

In 2001, the total market demand for PBDEs in the world was 67,390 metric tons. Technical decaBDE 

constituted about 56,100 metric tons (83%), while technical mixtures of octaBDEs and pentaBDEs were 

3,790 and 7,500 metric tons (6 and 11%) of this total, respectively (BSEF 2003). About 98% of the 

global demand for the technical pentaBDE mixture resides in North America (Hale et al. 2003). The 

Great Lakes Chemical Corporation recently announced that it is voluntarily phasing out production of 
PBBs and PBDEs 369 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

Table 7-1. Total Market Demand of PBDEs by Regions of the World in 2001a

 Americas Europe Asia Rest of world Total 

PentaBDE 7,100 150 150 100 7,500 

OctaBDE 1,500 610 1,500 180 3,790 

DecaBDE 24,500 7,600 23,000 1,050 56,100 

Source: BSEF 2003 

a

Data in metric tons 

decaBDE = commercial decabrominated diphenyl ether mixture; octaBDE = commercial octabrominated diphenyl 

ether mixture; pentaBDE = commercial pentabrominated diphenyl ether mixture 
PBBs and PBDEs 370 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

pentaDBEs and octaDBEs by the end of 2004. Since there are no other pentaDBE manufacturers in the 

United States, this decision functionally eliminated U.S. production of this compound (Tullo 2003). 

Table 7-2 lists the facilities in each state that manufacture or process technical decaBDE, the intended 

use, and the range of maximum amounts of technical decaBDE that are stored on-site. There are 

141 facilities that produce or process technical decaBDE in the United States (TRI01 2003). The data 

from the Toxics Release Inventory (TRI) listed in Table 7-2 should be used with caution, however, since 

only certain types of facilities were required to report. The TRI is not an exhaustive list. Facilities are 

only required to report to the TRI if they manufacture or process more than 25,000 pounds of a TRI listed 

chemical during the year, or otherwise use more than 10,000 pounds, and have the equivalent of more 

than 10 full-time employees. According to the EPA, TRI data have certain limitations. TRI data reflect 

releases and other waste management of chemicals, and not exposures of the public to those chemicals. 

TRI data alone are not sufficient to determine exposure or calculate potential adverse effects on human 

health and the environment. 

7.2 IMPORT/EXPORT 

Polybrominated Biphenyls. PBBs are no longer being imported or exported except possibly in small 

quantities for laboratory uses. PBBs have not been imported from other countries into the United States, 

except in finished products (Neufeld et al. 1977). The two companies that manufactured octa- and decabromobiphenyl in the United States between 1976 (0.805 million pounds) and 1978 exported all of their 

products to Europe (Neufeld et al. 1977). 

Polybrominated Diphenyl Ethers. No U.S. import or export data were located in the literature for 

PBDEs. In 2001, worldwide demands for technical penta-, octa-, and decabromodiphenyl ethers were 

7,500, 3,790, and 56,100 metric tons, respectively (see Table 7-1; BSEF 2003). 

7.3 USE 

Polybrominated Biphenyls. PBBs are no longer used in the United States. In the past, PBBs were used 

as additive flame retardants to suppress or delay combustion. Additive flame retardants are added to the 

polymer material, but are not chemically incorporated into the polymer matrix. Because PBBs are not 
PBBs and PBDEs 371 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

Table 7-2. Facilities that Produce, Process, or Use Decabromodiphenyl Ether 

Statea

Number of 

facilities 

Minimum amount on 

site in poundsb

Maximum amount on site in 

poundsb Activities and usesc

AL 1 10,000 99,999 7, 8 

AR 5 1,000 9,999,999 1, 4, 7, 8 

CA 5 1,000 999,999 7, 8 

CT 3 1,000 99,999 7, 8, 9, 14 

FL 2 100 999,999 7 

GA 5 100 99,999 2, 3, 5, 7, 8, 11, 12 

IL 2 10,000 999,999 7, 8 

IN 6 0 99,999 7, 8 

KY 2 10,000 99,999 7 

LA 1 100,000 999,999 12 

MA 12 1,000 99,999 7, 8 

MD 1 10,000 99,999 8 

MI 5 1,000 9,999,999 7, 8 

MN 3 10,000 99,999 7, 8 

MO 1 1,000 9,999 7 

MS 2 10,000 99,999 7, 8 

NC 13 1,000 99,999 7, 8 

NE 1 10,000 99,999 8 

NH 1 1,000 9,999 7 

NJ 4 10,000 999,999 7, 8 

NY 3 10,000 99,999 6, 7, 8 

OH 10 1,000 999,999 2, 3, 7, 8, 10 

PA 7 1,000 9,999,999 7, 8 

RI 1 1,000 9,999 7 

SC 9 1,000 999,999 7, 8 

TN 7 1,000 99,999 7 

TX 8 1,000 99,999 1, 2, 3, 7, 8, 12 

VA 5 10,000 999,999 7, 8 

VT 1 10,000 99,999 2, 5, 7, 12 

WA 1 10,000 99,999 8 

WI 2 1,000 9,999 7, 8 

Source: TRI01 2004 

a

Post office state abbreviations used 

b

Amounts on site reported by facilities in each state 

c

Activities/Uses:

1. Produce 

2. Import 

3. Onsite use/processing 

4. Sale/Distribution 

5. Byproduct 

6. Impurity 

7. Reactant 

8. Formulation Component 

9. Article Component 

10. Repackaging 

11. Chemical Processing Aid 

12. Manufacturing Aid 

13. Ancillary/Other Uses 

14. Process Impurity 
PBBs and PBDEs 372 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

chemically bound to the polymer matrix, they may migrate out of the matrix with time (WHO 1994b). 

PBB applications were almost exclusively limited to a particular thermoplastic (arylonitrile-butadienestyrene, ABS) used in electronic equipment housings (Hardy 2002b). Prior to termination of production, 

hexabromobiphenyl was used as a fire retardant mainly in thermoplastics for constructing business 

machine housings and in industrial (e.g., motor housing), and electrical (e.g., radio and TV parts) 

products. Smaller amounts were used as a fire retardant in coating and lacquers, and in polyurethane 

foam for auto upholstery (Neufeld et al. 1977). PBDEs and other flame retardants replaced hexabromobiphenyl after its voluntary ban in the late 1970s. Octabromobiphenyl and decabromobiphenyl were 

never used in the United States, probably because the hexabromobiphenyl was less expensive and equally 

effective as a fire retardant (Neufeld et al. 1977). 

Polybrominated Diphenyl Ethers. PBDEs are used as additive flame retardants in thermoplastics. 

Additive flame retardants are physically combined with the polymer material being treated rather than 

chemically combined (as in reactive flame retardants). This means that there is a possibility that the 

flame retardant may diffuse out of the treated material to some extent. PBDEs are used in different resins, 

polymers, and substrates at levels ranging from 5 to 30% by weight (EU 2001). As of August 15, 2004, 

no products containing more than 0.1% penta- or octoBDE by mass can be sold in the European Union 

(EU 2003b). 

The commercial pentaBDE product is used predominantly (95–98%) for flame retardant purposes as an 

additive in consumer products manufactured by the furniture industry (ENVIRON 2003a). It is used 

almost exclusively for flame retard flexible polyurethane foam (FPUF), which is used in bed mattresses 

and cushioning in upholstered products. The commercial pentaBDE is typically used in FPUF as an 

additive mixture with aromatic phosphate esters (e.g., mixture of 75% pentaBDE and 25% aromatic 

phosphate esters). Mattress FPUF contains approximately 2–3% flame retardant mixture and cushion 

FPUF contains 3–5% flame retardant mixture (ENVIRON 2003a). Scrap materials from both industries 

have been used as padding beneath carpets, and as a result, carpet padding likely contains 3–5% flame 

retardant mixture. However, not all of the FPUF found in cushion, mattress, and carpet padding products 

are treated with commercial pentaBDE. Approximately 7.5% of the more than 2.1 billion pounds of 

FPUF produced annually in the United States uses the commercial pentaBDE product as a flame retardant 

additive (ENVIRON 2003a). The majority of FPUF products treated with the commercial pentaBDE 

product are sold in California, the only state requiring by law that upholstered products achieve a 

prescribed level of ignition resistance (ENVIRON 2003a). A small percentage of pentaBDE is used in 

commercial adhesive products. Historical uses of commercial pentaBDE included coatings for specialty 
PBBs and PBDEs 373 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

textiles, printed circuit board components, hydraulic and oilfield completion fluids, and rubber products. 

However, all of these uses have been discontinued in recent years. In the past, automotive and airplane 

seating cushions contained FPUF with commercial pentaBDE. However, this use was discontinued in the 

early 1990s. Prior to approximately 1990, the commercial pentaBDE product may have been used in 

small quantities as a flame retardant in specialty fire-resistant clothing using polyurethane treatment and 

in polyurethane coatings in carpets (ENVIRON 2003a). With the exception of small quantities used in 

rigid polyurethane elastomers for instrument casings, the commercial pentaBDE product is not used in 

electronic equipment. However, it is possible that electronic equipment containing pentaBDE produced 

in other countries (principally Asian) could find its way into the United States (EU 2001). The 

commercial pentaBDE product is not used in acrylonitrile-butadiene-styrene plastics or electronics 

equipment (ENVIRON 2003a). Other applications of pentaBDE may include use as an additive flame 

retardant for flexible polyester polyurethane foam, rigid polyurethane, rigid polyurethane foams, epoxies, 

laminates, adhesives, coatings, unsaturated polyesters, and flexible polyvinyl chloride (PVC) compounds 

(Great Lakes Chemical Corporation 2004a). 

The commercial octaBDE is used by the plastics industry as an additive flame retardant for manufactured 

products. It is used almost exclusively to flame retard acrylonitrile-butadiene-styrene (ABS) terpolymers 

used in computer casings and monitors (ENVIRON 2003b). In the European Union, approximately 95% 

of the total commercial octaBDE product sold to the electronics and plastics industries is used in ABS 

(EU 2003a). Although data are not available in the United States, similar volumes are expected 

(ENVIRON 2003b). The commercial octaBDE product is used in ABS products at 12–18% weight 

loadings of flame retardant. OctaBDE is always used as a flame retardant in conjunction with antimony 

trioxide. Other minor uses for octaBDE, accounting for the remaining 5% use, include high impact 

polystyrene (HIPS), polybutylene terephthalate (PBT), and polyamide polymers (EU 2003a). Other 

possible applications of octaBDE include use as additive flame retardant in polycarbonate, phenolformaldehyde resins, unsaturated polyesters (EU 2003a), nylon, thermoplastic elastomers, polyolefins, 

adhesives, and coatings (Great Lakes Chemical Corporation 2004b). 

The commercial decaBDE product is an additive flame retardant used in a variety of polymer 

applications. Industry information indicates that decaBDE is used at loadings of 10–15% weight in 

polymers and is always used in conjunction with antimony trioxide (EU 2002). The major application for 

decaBDE is in high impact polystyrene (HIPS), which is used in the television industry for cabinet backs. 

It is also used for a large number of other polymers with end-uses in electrical and electronic equipment 

(e.g., computers, connectors, electrical boxes, wire and cable, etc.). Examples include polypropylene (for 
PBBs and PBDEs 374 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

electronics), acetate copolymers (ethylene-vinyl acetate [EVA] and other copolymers for wire and cable), 

ethyelene-propylene-diene terpolymer (EPDM) and thermoplastic elastomers (for wire and cable), and 

polyester resins (for electronics). Other minor uses include styrenic rubbers, polycarbonates, polyamides, 

and terphthalates, and small amounts are reported to be used in hot-melt adhesives (EU 2002). Other 

possible applications of decaBDE are as an additive flame retardant for low-density polyethylene (LDPE), 

unsaturated polyester, epoxy, coatings, adhesive systems, and backcoatings for fabrics (Great Lakes 

Chemical Corporation 2004c). Upholstered furniture in commercial settings in the United States is 

required to meet federal flammability standards and may utilize upholstery textiles that are flame retarded 

with a backcoating containing decaBDE at 5 mg/m2 (BFRIP 2002). 

Currently, technical decaBDE is the most widely used PBDE flame retardant worldwide, followed by 

technical pentaDBE. About 95% of technical pentaBDE is used in the manufacture of flexible 

polyurethane foam used as cushioning in upholstery. Flexible polyurethane foam is also used in foambased laminated automotive applications, for domestic furniture, and in foam-based packaging (EU 

2001). Technical octaBDE is mainly used in the preparation of acrylonitrile-butadiene-styrene 

terpolymer (ABS), which is used in the manufacture of computer and business equipment housings. 

Technical octaBDE is also used in adhesives and coatings applications (WHO 1994a). Technical 

decaBDE is primarily used in combination with antimony trioxide in high impact polystyrene (HIPS) 

applications, such as electronic enclosures (e.g., television cabinets). Other minor uses of technical 

decaBDE are in textile applications to flame retard upholstery fabric (e.g., polyester fiber additives and 

coatings for automobile fabric, tarpaulins, and tents) (WHO 1994a). 

7.4 DISPOSAL 

Polybrominated Biphenyls. PBBs are no longer commercially produced in the United States. In the past, 

an estimated 0.0046 pounds have been lost to sewers for every 1,000,000 pounds of PBBs produced at 

manufacturing sites (Neufeld et al. 1977). The Michigan Chemical Corporation discharged an estimated 

0.25 pounds of PBBs/day to the Pine River as effluent (Di Carlo et al. 1978). The Michigan Chemical 

Corporation estimated that the solid waste generated during the manufacture of FireMaster BP-6 was 5% 

of the FireMaster BP-6 and FireMaster FF-1 produced (Di Carlo et al. 1978). Since Michigan Chemical 

Corporation produced ≈11.8 million pounds of FireMaster BP-6 and FireMaster FF-1 from 1970 to 1974 

(Di Carlo et al. 1978), solid wastes containing a total of 590,000 pounds of PBBs would have been sent to 

disposal. About one-half of this waste was deposited in the Gratiot County landfill in St. Louis, Michigan 

(Di Carlo et al. 1978), and the rest was possibly landfilled at other locations. Contaminated animal 
PBBs and PBDEs 375 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

carcasses, poultry and eggs, animal feed, butter, cheese, and other milk products following the Michigan 

agriculture contamination episode were disposed of in a sanitary landfill in Cadillac, Michigan (Dunckel 

1975). 

Approximately 11.8 million pounds of hexabromobiphenyl were used in commercial and consumer 

products in the United States, most in the production of plastic products with an estimated use life of 5–

10 years (Neufeld et al. 1977). Since the cessation of production, all of these products, such as TV 

cabinet and business machine housings, must have been disposed of by land filling or incineration 

(Neufeld et al. 1977). The formation of polybrominated dioxins (PBDDs) and polybrominated dibenzofurans (PBDFs) during the incineration of plastics containing PBBs remains a distinct possibility (Luijk 

and Govers 1992; O'Keefe 1978). 

Polybrominated Diphenyl Ethers. PBDEs are currently used as flame retardants in a wide range of 

consumer products (see Section 7.3). In the United States, waste disposal of PBDE-containing consumer 

products is described as transfers to disposal (landfill), recycling, energy recovery (incineration), or 

publicly owned treatment works (POTWs) (Darnerud et al. 2001). No other information was located on 

the past or present volumes of PBDE-containing consumer products disposed of by each method of waste 

transfer. 

Landfill disposal of plastic consumables containing pentaBDE (e.g., polyurethane foams), octaBDE (e.g., 

computer monitors) and decaBDE (e.g., televisions) to landfills is likely to increase in the United States 

due to their limited useful lifespan. Given that all PBDEs have low water solubilities (see Table 6-6), the 

potential for leaching of PBDE from landfills appears to be small (EU 2002). Well-designed landfills will 

include measures to minimize leaching and those measures would also be effective in minimizing the 

leaching of any PBDE particulates present (EU 2002). 

PBDEs in sewage are disposed of in publicly owned treatment works (POTWs), as indicated by analysis 

of sewage sludge from various countries. Tetra- and penta-BDEs have been detected in sewage sludge 

from POTWs in the United States with the composition of PBDEs in these biosolids closely resembling 

the DE-71 (i.e., pentaBDE) commercial formulation (La Guardia et al. 2000). Hale et al. (2001b) 

reported that the practice of land application of sewage sludges may introduce significant amounts of the 

pentaBDE commercial mixture into the environment, although levels have not yet been quantified. 
PBBs and PBDEs 376 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

Incineration of waste materials containing PBDEs is thought to be a potential source of PBDFs and/or 

PBDDs. The formation of PBDFs/PBDDs as a result of uncontrolled landfill fires is also a possibility, 

although no data are available on the scale of this source. The results of pyrolysis experiments showed 

that PBDEs can form PBDFs and PBDDs (in much smaller quantities) under a wide range of heating 

conditions. If chlorine is present, mixed halogenated furans/dioxins can be formed (Oberg et al. 1987; 

Zier et al. 1991). Unless sufficiently high temperatures and long residence times are maintained, 

PBDFs/PBDDs can be generated during the incineration of products containing PBDEs. When heavy 

metals are present, the concentration of PBDDs and PBDFs are higher than when no metals are present. 

Sakai et al. (2001a) measured residues of PBDFs/PBDDs in effluents from a municipal incineration plant 

burning domestic waste materials. Flue gases, fly ash, and bottom ash reportedly contained 

PBDFs/PBDDs at concentration ranges of 0.28–3.3 ng/N m3

, 0.082–13 ng/g, and 0.0058–27 ng/g, 

respectively. However, modern, properly operated municipal waste incineration should not emit 

significant quantities of PBDFs/PBDDs, regardless of the composition of municipal waste (WHO 1994a). 

In the United States, waste disposal of industrial by-products containing PBDEs may also be described as 

transfers to disposal (landfill), recycling, energy recovery (incineration), industrial treatment works, or 

POTWs. The types of waste transfer may be different for manufacturing versus processing sectors, and 

also from within different types of processing. Waste disposal from manufacturing processes is 

predominantly to secure chemical landfills (e.g., those built with liners and leachate collection). For 

example, solid wastes (e.g., filter cakes) from the commercial production of technical octaBDE are 

disposed of in hazardous waste landfills as these wastes may contain toluene from the production process 

(EPA 1995). Plastic processors typically transfer most waste to disposal (landfill), recycling, energy 

recovery (incineration), and industrial treatment works, while minimal releases are to POTWs. In 

contrast, textile processors typically transfer most waste to POTWs. This difference in waste transfers 

between the plastic and textile sectors is because textile processors use water in their processing operation 

and other processors (e.g., processors of plastic) do not. 

Recycling of plastic materials containing PBDEs is a common practice in industry. It has been 

demonstrated that decaBDE-containing resins can be successively recycled without generation of 

PBDDs/PBDFs (Brenner and Knies 1990; Donnelly et al. 1987); McAllister et al. 1990). For example, 

virgin high impact polystyrene (HIPS) resins (containing antimony trioxide [Sb2O3] and decaBDE) and 

repeatedly ground and injected molded (e.g., “recycled”) HIPS/decaBDE/Sb2O3 resins both met the 

requirements of the German Chemicals Banning Ordinance with respect to 2,3,7,8-substituted 

PBDD/PCDF congeners. These resins were at least 1 order of magnitude below the regulated limit values 
PBBs and PBDEs 377 

7. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

for PBDDs/PCDFs (1 ppb for the sum of four congeners, 5 ppb for the sum of all eight regulated 

congeners). 

PBBs and PBDEs 379 

8. POTENTIAL FOR HUMAN EXPOSURE 

8.1 OVERVIEW 

Polybrominated Biphenyls. PBBs have been identified in at least 9 of the 1,647 hazardous waste sites 

that have been proposed for inclusion on the EPA National Priorities List (NPL) (HazDat 2004). 

However, the number of sites evaluated for PBBs is not known. The frequency of these sites can be seen 

in Figure 8-1. Of these sites, all are located within the United States and none are located in the 

Commonwealth of Puerto Rico (not shown). 

The production of PBBs in the United States ceased in 1979 (IARC 1986). In the past, PBBs were 

released to the environment during the manufacture of these compounds and disposal of commercial and 

consumer products containing these compounds (Hesse and Powers 1978; Neufeld et al. 1977). One of 

the significant sources of environmental contamination occurred as a result of the accidental mixup of 

FireMaster BP-6 with cattle feed in a number of farms in the lower peninsula in Michigan (see 

Section 5.2 for additional details concerning this incident). By June 1975, 412 farms had been 

quarantined. Disposal of contaminated feed, animal carcasses (poultry, dairy cattle, swine), and animal 

products (dairy, meat, eggs) contributed to environmental contamination (Dunckel 1975; Kay 1977). No 

information was located on the current levels of contamination at these locations. 

PBBs can exist as 209 different congeners, but only about 42 have been synthesized (Sundstrom et al. 

1976b). Environmental contamination of PBBs is likely to have occurred mainly from the two 

commercial products, FireMaster BP-6 and FireMaster FF-1. The principal component in both of these 

commercial products was 2,2',4,4',5,5'-hexabromobiphenyl or BB-153 (Robertson et al. 1983b). 

PBBs are strongly adsorbed to soil and sediment (Filonow et al. 1976; Hesse and Powers 1978) and 

usually persist in the environment (Jacobs et al. 1978). Adsorption of PBBs generally increases as 

bromination of the PBBs and organic carbon content of soil and sediment increase (Filonow et al. 1976; 

Griffin and Chou 1981a, 1981b). As a result, the leaching of commercial mixtures of PBBs from soil is 

slow. Leaching studies with four Michigan soils mixed with 100 mg/kg 2,2',4,4',5,5'-hexabromobiphenyl 

showed that <0.6% of the compound leached through soils after a 19-day period. Leachate quantities in 

this study were equivalent to 20 times the average annual rainfall in Michigan (Filonow et al. 1976). The 

PBBs in commercial mixtures resist both chemical and biological degradation (Jacobs et al. 1978; 
PBBs and PBDEs 380 

8. POTENTIAL FOR HUMAN EXPOSURE 

Figure 8-1. Frequency of NPL Sites with PBB Contamination 

1

4

Derived from HazDat 2004

Frequency of

NPL Sites
PBBs and PBDEs 381 

8. POTENTIAL FOR HUMAN EXPOSURE 

Kawasaki 1980; Shelton and Tiedje 1981), although biotic debromination to lower brominated products 

may occur in anaerobic zones of contaminated sediment and soil (Morris et al. 1992). 

PBBs with six or fewer bromine substitutions bioconcentrate in aquatic organisms such as fish, but the 

octabromo- and decabromobiphenyls do not bioconcentrate significantly in fish (Gobas et al. 1989; Norris 

et al. 1973; Opperhuizen et al. 1985; Veith et al. 1979; Zitko 1979; Zitko and Hutzinger 1976). Orchard 

grass, alfalfa, corn, and tops of carrots grown in soil contaminated with PBBs showed no uptake of PBBs, 

and only minor uptake occurred on carrot roots (Jacobs et al. 1976, 1978). Although PBBs were detected 

in fish-eating birds and predatory animals that had consumed PBB-contaminated food (Heinz et al. 1983, 

1985), the biomagnification potential of PBBs in predators resulting from such consumption remains 

unknown. 

PBBs were detected in air, water, sediment, and soil in the vicinity of the manufacturing plants and in 

groundwater from a landfill site (DeCarlo 1979; Hesse and Powers 1978; Shah 1978). PBBs were also 

detected in soil near the contaminated farms in Lower Michigan (Fries and Jacobs 1980). The 

distribution of PBBs was limited to the environment in the vicinity of production sites and the 

contaminated farm sites. Recent studies have identified PBBs in marine mammals from coastal seas and 

the Atlantic Ocean (de Boer et al. 1998). Data regarding the current levels of PBBs in ambient air, 

drinking water, or food were not located. 

No estimate on PBB intake by the general population from air, water, and food was located in the 

literature. Current intake of PBBs for the general population is expected to be zero or very small. 

Populations near the contaminated farms in Lower Michigan may still have low exposures from air, 

water, and food. The level of PBBs in human tissue and body fluids in the exposed population of 

Michigan has been extensively studied (Brilliant et al. 1978; Cordle et al. 1978; Eyster et al. 1983; 

Humphrey and Hayner 1975; Lambert et al. 1990; Landrigan et al. 1979; Wolff et al. 1979a, 1982). The 

finding that PBBs are stored in fatty tissues of the human body and are very slowly excreted (Eyster et al. 

1983) indicates a slow decline in the body burden for exposed individuals. 

Polybrominated Biphenyl Ethers. The widespread use of PBDEs over the past 30 years has resulted in 

the presence of some lower-brominated congeners in the environment (e.g., 2,2’,4,4’-tetrabromodiphenyl 

ether [BDE 47]). However, highly brominated congeners (e.g., decaBDE) are typically detected only 

near point sources. PBDEs are released into the environment from their manufacture and use as additive 

flame retardants in thermoplastics in a wide range of products (WHO 1994a). PBDEs containing waste 
PBBs and PBDEs 382 

8. POTENTIAL FOR HUMAN EXPOSURE 

may be incinerated as municipal waste, deposited in landfills, discharged to municipal sewage-treatment 

plants, or emitted directly to the atmosphere as particulates (Darnerud et al. 2001). In the future, the 

disposal of plastic consumables containing PBDEs is likely to increase in the United States (NSC 1999). 

PBDEs are strongly adsorbed to soil and sediment and persist in the environment. Adsorption of PBDEs 

generally increases as bromination of PBDEs and organic carbon content of soil and sediment increase. 

As a result, most PDBEs have little or no mobility in soil and are not expected to leach (e.g., into 

groundwater). Lower BDE homologs (e.g., tri- and tetraBDE), which may exist partially in the vapor 

phase, have the potential for long-range transport in the atmosphere. The detection of lower brominated 

PBDEs (e.g., 2,2’,4,4’-tetraBDE [BDE 47]) in remote regions of the world suggests that long-range 

transport of these congeners is occurring. Higher BDE homologs (e.g., decaBDE) will primarily exist 

near point sources. Biodegradation will not be significant for PBDEs, but under certain conditions, some 

PBDEs compounds (e.g., decaBDE) may degrade by direct photolysis to form lower-brominated 

congeners. However, determining the rate and extent of degradation processes (e.g., biodegradation and 

photolysis) for PBDEs, such as decaBDE and pentaBDE commercial mixtures, is still an active area of 

research. 

Monitoring studies indicate that lower brominated PBDEs, for example, 2,2’,4,4’-tetrabromodiphenyl 

ether (BDE 47) and 2,2’4,4’,5-pentabromodiphenyl ether (BDE 99), are transported globally. 

Atmospheric, water, and biota levels of PBDEs also tend to be dominated by lower brominated congeners 

(e.g., BDE 47, BDE 99). Environmental concentrations of lower brominated PBDEs (i.e., tetraBDE and 

pentaBDE) appear to be leveling off in Europe, but appear to be increasing in certain areas in Canada and 

the United States, although data are too sparse to make broad statements regarding trends. In the 

environment, higher brominated commercial mixtures (e.g., decaBDE) are concentrated in soils and 

sediment near industrial point sources. 

Studies of the biota indicate that lower brominated congeners (e.g., 2,2’,4,4’-tetraBDE [BDE 47]) are 

being preferentially bioconcentrated. Lower brominated diphenyl ether (e.g., tetra- and penta-) 

concentrations increase with respect to trophic level; thus, organisms that reside higher on food chains 

tend to have higher concentrations of these brominated diphenyl ethers. Body-burden data indicate that 

the general population is exposed to low levels of lower brominated (e.g., tetra- and penta-) BDEs. In 

general, decaBDE is not found in body burden data. 
PBBs and PBDEs 383 

8. POTENTIAL FOR HUMAN EXPOSURE 

Humans are primarily exposed to lower brominated (e.g., tetra- and pentaBDEs) BDEs by inhalation of 

ambient or contaminated air and ingestion of contaminated food. Levels of PBDEs in body tissues and 

fluids from individuals living in the United States have recently been determined. Most studies indicate 

that levels of lower brominated BDEs (e.g., 2,2’,4,4’-tetraBDE [BDE 47]) in body fluids and tissues are a 

factor of 10–100-fold higher for individuals living in the United States compared to individuals living in 

other regions of the world (e.g., Europe). In general, decaBDE is not detectable in body fluids and 

tissues. Occupational exposure to PBDEs occurs primarily by inhalation of air containing PBDEs. 

PBDEs have not been identified in any of the 1,647 hazardous waste sites that have been proposed for 

inclusion on the EPA National Priorities List (NPL) (HazDat 2004). However, the number of sites 

evaluated for PBDEs is not known. 

8.2 RELEASES TO THE ENVIRONMENT 

Polybrominated Biphenyls. The production of PBBs in the United States ceased in 1979 (IARC 1986). 

In the past, PBBs were released to the environment during the manufacture of these compounds and 

disposal of commercial and consumer products containing these compounds (Hesse and Powers 1978; 

Neufeld et al. 1977). One of the significant sources of environmental contamination occurred as a result 

of the accidental mixup of FireMaster BP-6 with cattle feed in a number of farms in the lower peninsula 

in Michigan (see Section 5.2 for additional details concerning this incident). By June 1975, 412 farms 

had been quarantined. Disposal of contaminated feed, animal carcasses (poultry, cattle, swine), and 

animal products (meat, milk, eggs) contributed to environmental contamination (Dunckel 1975; Kay 

1977). 

Polybrominated Diphenyl Ethers. The widespread use of PBDEs over the past 30 years has resulted in 

the presence of lower brominated diphenyl ether congeners, for example, 2,2’,4,4’-tetrabromodiphenyl 

ether (BDE 47) and 2,2’4,4’,5-pentabromodiphenyl ether (BDE 99), in the environment. Higher 

brominated congeners (e.g., octabromodiphenyl ether and decabromodiphenyl) tend to concentrate near 

point sources (Wania and Dugani 2002). The commercial production of PBDEs began in late 1970s 

(WHO 1994a). In 2001, the total market demand for PBDEs in the Americas was 33,100 metric tons 

(BSEF 2003). Technical decabromodiphenyl ether constituted about 24,500 metric tons (74%), while 

technical mixtures of octa- and pentabromodiphenyl ethers were 1,500 and 7,100 metric tons (4 and 22% 

of this total), respectively (BSEF 2003). PBDEs may be released into the environment from their 

manufacture and use in a wide range of consumer products (WHO 1994a). PBDEs are used as additive 
PBBs and PBDEs 384 

8. POTENTIAL FOR HUMAN EXPOSURE 

flame retardants in thermoplastics. Additive flame retardants are physically, rather than chemically, 

combined with polymers. Thus, there is a possibility that some PBDEs congeners (e.g., lower brominated 

congeners) may diffuse out of the treated materials to some extent (EU 2001). Waste from products 

containing PBDEs may be either incinerated as municipal waste, deposited in landfills, or discharged to 

municipal sewage-treatment plants (Darnerud et al. 2001). In the future, the disposal of plastic 

consumables containing PBDEs to landfills is likely to increase in the United States and elsewhere in the 

world. 

The widespread use of pentaBDE over the past 30 years has resulted in the rapid increase in 

concentrations of lower brominated congeners, for example, 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47) 

and 2,2’4,4’,5-pentabromodiphenyl ether (BDE 99), and their presence in the environment. No 

quantitative information was located on the releases of the commercial pentaBDE product from its 

production and use in the United States. The commercial pentaBDE product is used predominantly (95–

98%) as an additive flame retardant in flexible polyurethane foam (FPUF), which has applications as 

cushioning in bed mattresses and upholstered products (see Section 7.3). No quantitative information was 

located on the releases of pentaBDE from the production of flexible polyurethane foams. However, the 

main source of release for liquid flame retardant additives (e.g., pentaBDE) is associated with the handing 

of the raw material (e.g., splashes and spills) prior to the foaming process, where releases are to waste 

water (EU 2001). There is a potential release to air during the curing phase since the polyurethane foam 

is heated at elevated temperatures (e.g., up to 160 °C) for several hours. Foam scrap will be disposed of 

to a landfill or possibility incinerated (EU 2001). Since pentaBDE is an additive flame retardant, it may 

be subject to volatilization or leaching from the polymer matrix during the lifetime of the use of an article. 

Losses of foam particles containing the substance (e.g., due to abrasion) may also occur. However, 

pentaBDE has a very low vapor pressure (see Table 6-6) and volatilization losses from polyurethane foam 

would be expected to be low. Given that the major use of pentaBDE is as a component of polyurethane 

foam for furniture/seating/automobile use, the potential for leaching of pentaBDE during use will be 

minimal. This is because, although it is likely that foam coverings may be washed during the lifetime of 

use, it is very unlikely that the foam cushioning containing pentaBDE will be washed as well (EU 2001). 

Residual ‘waste’ pentaBDE in the environment will be particles of polymer (foam) products that contain 

pentaBDE (e.g., 2,2’4,4’-tetraBDE). Polyurethane foam has an open cell structure, which presents a large 

surface area to the environment and therefore, the potential for release is likely to be greater than the hard 

dense plastics where octaBDE and decaBDE are used. In addition, this foam may become friable and 

crumble with age. Thus, small particles thereby released could move into the environment and disperse 

its components (i.e., commercial pentaBDE mixture) (BFRIP 2002). These form particles are released to 
PBBs and PBDEs 385 

8. POTENTIAL FOR HUMAN EXPOSURE 

the urban/industrial soil compartment, but may also end up in sediment or air (EU 2001). Movement of 

polymer (foam) particles containing pentaBDE within the landfill could provide a transport mechanism 

leading to entry into leachate water or groundwater. However, well-designed landfills, in general, already 

include measures to minimize leaching and these measures would also be effective in preventing any 

leaching of pentaBDE as well (EU 2001). 

Commercial octaBDE is used almost exclusively to flame retard acrylonitrile-butadiene-styrene (ABS) 

terpolymers used in computer casings and monitors (ENVIRON 2003b). No quantitative information was 

available on emissions of octaBDE from production operations. The major sources of air emissions are 

thought to be as a result of grinding and bagging operations. The emission to air of octaBDE vapor from 

production can be considered to be negligible (EU 2003a). The most likely way in which octaBDE may 

reach water from its production is due to washing out of equipment. The major source of octaBDE waste 

is from filter waste and reject material (EU 2003a). Much of the loss from polymer applications (e.g., 

acrylonitrile-butadiene-styrene) is likely to be in the form of dust. It is expected that much of this dust 

will be collected for reuse or disposed of to landfill/incineration. Some of this may end up in waste water 

as a result of cleaning floors and equipment (EU 2003a). Given that the major use of plastics containing 

octaBDE appears to be in electrical applications and that the substance has very low water solubility, the 

potential for leaching of octaBDE from the products during uses appears to be small. Waste remaining in 

the environment can be considered to be particles (or dust) of polymer product, or dust generated from 

polymer products, that contain octaBDE. These particles are primarily released to urban/industrial soil, 

but may also end up in sediment or air. Plastics containing octaBDE will usually be disposed of either to 

landfills or by incineration. No experiments have been carried out on the leachability of octaBDE from 

polymers in landfills. However, octaBDE is not expected to leach to a significant extent from polymers, 

unless the polymer itself undergoes some form of degradation, thus releasing octaBDE. It is expected that 

emissions of octaBDE from incineration processes will be near zero (EU 2003a). 

Commercial decaBDE product is an additive flame retardant used in a variety of polymer applications 

(EU 2002). The major application for decaBDE is in high-impact polystyrene (HIPS), which is used in 

the television industry for cabinet backs. No quantitative information is available on emissions of 

decaBDE from production operations. Since the major use of plastics containing decaBDE is in 

electrical/electronic applications and the substance has very low water solubility, the potential for 

leaching of decaBDE from plastic products during use appears to be small (EU 2003a). A study of the 

leaching of decaBDE (DOW FR-300-BA; 77.4% deca-, 21.8% nona-, and 0.8% octaBDE) from pellets of 

acrylonitrile-butadiene-styrene (ABS) polymer and polystyrene (both containing 10% decaBDE) was 
PBBs and PBDEs 386 

8. POTENTIAL FOR HUMAN EXPOSURE 

undertaken. The lack of increase in bromine concentration with time was observed. In static extractions 

of ABS containing 4.25% decaBDE (Dow FR-300-BA) with water, acetic acid, and cottonseed oil at 

elevated temperatures, little or no leaching of the decaBDE was evident. No decaBDE was detected in 

the water and acetic acid and only 0.03% of the total decaBDE was extracted by cottonseed oil over 

7 days at elevated temperatures (EU 2002). 

Releases of decaBDEs are required to be reported under the Superfund Amendments and Reauthorization 

Act (SARA) Section 313; consequently, data are available for this compound in the Toxics Release 

Inventory (TRI) (EPA 1995). According to the TRI, a total of 649,541 pounds (294,627 kg) of decaBDE 

were released to the environment in 2001. In addition, an estimated 787,927 pounds (357,398 kg) were 

transferred off-site, including to publicly-owned treatment works (POTWs) (TRI01 2004). The TRI data 

should be used with caution because only certain types of facilities are required to report. This is not an 

exhaustive list. Facilities are only required to report to the TRI if they manufacture or process more than 

25,000 pounds of a TRI listed chemical during the year, or otherwise use more than 10,000 pounds, and 

have the equivalent of more than 10 full-time employees. According to the EPA, TRI data have certain 

limitations. TRI data reflect releases and other waste management of chemicals and not exposures of the 

public to those chemicals. TRI data alone are not sufficient to determine exposure or calculate potential 

adverse effects on human health and the environment. 

On-site land releases dominate decaBDE releases on-site from its manufacture and use, and are 

predominantly associated with disposal of decaBDE in either a manufacturer’s on-site landfill in 

Arkansas or in a commercial chemical landfill in Louisiana. On-site air and water releases make up only 

a small fraction of total on-site releases. For example, on-site air emissions in 2000 were ~6.6% of the 

total; on-site surface water discharges also make up only a small fraction of the total (~0.2% of the 

2000 total) and are nearly all associated with operations that apply decaBDE to upholstery textiles. 

DecaBDE releases off-site for disposal from its manufacture and use are typically larger than those onsite to land; decaBDE-waste transfers for further waste management are dominated by transfers to 

POTWs from the textile industry, although a significant rise in recycling has occurred in recent years. 

The plastics industry released minimal amounts to POTWs. The largest releases of decaBDE or 

decaBDE-waste to water occur in Maryland, North Carolina, and South Carolina and result from its use in 

textiles (TRI01 2004). 
PBBs and PBDEs 387 

8. POTENTIAL FOR HUMAN EXPOSURE 

8.2.1 Air 

Polybrominated Biphenyls. In the past, PBBs were released into the air during the manufacture of these 

compounds in three areas: through the vents of the hydrogen bromide recovery system, from the 

centrifugation area for recovering PBBs from slurries produced by bromination, and from the drying, 

pulverizing, and bagging area of the finished product (Di Carlo et al. 1978). An estimated 

0.07 pounds/million pounds of the PBBs produced were lost from the hydrogen bromide-recovery vent 

(Di Carlo et al. 1978). No data are available for the air pollution factor (amount released/million pounds 

produced) at the centrifugation site. The concentrations of FireMaster BP-6 in the Michigan Chemical 

Corporation bagging area were 0.016–0.032 mg/L of air during the bagging operation and 0.003 mg/L of 

air after the completion of bagging (Di Carlo et al. 1978). In 1977, the maximum air losses of PBBs at 

production sites were estimated to total 1,125 pounds of PBBs for every 1 million pounds of PBBs 

produced (Di Carlo et al. 1978). 

Another process that could release lower levels of brominated biphenyls in the air is the incineration of 

PBBs. Pyrolysis of hexabromobiphenyl in the absence and presence of air has produced small amounts of 

lower brominated biphenyls (Thoma and Hutzinger 1987). No data are available on the importance of 

this source for the release of PBBs in the air during the incineration of PBBs. However, since the vast 

majority of products containing PBBs are expected to be out of circulation after more than 25 years since 

the voluntary ban, incineration will not be a significant source of PBBs to air. 

PBBs have been identified in 1 air sample, collected from 1,647 NPL hazardous waste sites, where they 

were detected in some environmental media (HazDat 2004). 

Polybrominated Diphenyl Ethers. The widespread use of pentaBDE technical mixtures over the past 

30 years has resulted in the increasing concentrations of lower brominated congeners in air, e.g., 

2,2’,4,4’-tetrabromodiphenyl ether (BDE 47) and 2,2’4,4’,5-pentabromodiphenyl ether (BDE 99). No 

quantitative information was located on the releases of the pentaBPE technical mixtures to the atmosphere 

from its production and use. However, the release of pentaBDE technical mixtures to air has the potential 

to occur during the curing phase, since the polyurethane foam is at elevated temperatures (e.g., up to 

160 °C) for several hours during this phase. Since pentaBDE technical mixtures are additive flame 

retardants, they may be subject to volatilization or leaching from the polymer matrix during the lifetime 

of the use of the foam article. Losses of foam particles containing the substance (e.g., due to abrasion) 

may also occur. However, most congeners in pentaBDE technical mixtures have very low vapor 

pressures (see Table 6-6) and therefore, losses from polyurethane foam due to volatilization would be 
PBBs and PBDEs 388 

8. POTENTIAL FOR HUMAN EXPOSURE 

expected to be low. Migration of pentaBDE technical mixtures from consumer products may be a 

significant diffuse source of lower brominated congeners of pentaBDE technical mixtures to the 

atmosphere. Although no studies were found that determined the migration rate of pentaBDE technical 

mixtures from polymers into the air, estimates have been made. The estimated migration rate for 

pentaBDE technical mixtures is 0.39% per year (Danish EPA 1999). 

In the regions surrounding metal-recycling plants in Taiwan and Japan, a variety of tri-, tetra-, and 

hexabromodiphenyl ethers (BDEs) were measured in air (Watanabe et al. 1992). The concentrations in 

Taiwan and Japan were 23–53 and 7.1–21 pg/m3

, respectively. 

No quantitative information is available on emissions of octaBDE technical mixtures to the atmosphere 

from production operations. The major sources of air emissions of octaBDE technical mixtures are 

thought to be a result of grinding and bagging operations. However, the emission to air of vapors from 

production of octaBDE technical mixtures may be considered to be negligible (EU 2003a). OctaBDE 

technical mixtures are used almost exclusively to flame retard ABS terpolymers used in computer casings 

and monitors (ENVIRON 2003b). No quantitative information was available on emissions of octaBDE 

technical mixtures to the atmosphere from ABS applications. However, much of the loss from polymer 

applications is likely to be in the form of dust. Losses of octaBDE technical mixture powders (particle 

sizes >40 μm) have been estimated as 0.21% during handling of raw materials. These losses will initially 

be to the atmosphere, but it is expected that this dust will rapidly settle and may be disposed of to 

landfills. OctaBDE technical mixtures in the compounding stage are also susceptible to dust generation. 

However, losses at this stage are thought to be lower than during the handling of raw materials (EU 

2003a). Plastics containing octaBDE technical mixtures will usually be disposed of either to landfill or 

by incineration. It is expected that emissions of PBDE congeners resulting from incineration of plastics 

containing octaBDE technical mixtures will be negligible (EU 2003a). 

No quantitative information is available on emissions of decaBDE technical mixtures to the atmosphere 

from production operations. In 1979, decaBDE was found in the atmosphere as particulate matter in the 

vicinity of plants manufacturing brominated flame retardants (Zweidinger et al. 1979); the concentration 

of decaBDE ranged from not detected to 72 ng/m3

. The estimated releases of decaBDE vapor to air 

during production are low, typically 1.1x10-5 mg/metric tons of production (EU 2002). The major source 

of air emissions is thought to be a result of grinding and bagging operations. Losses of powders during 

the handling of raw materials will initially be to the atmosphere, but it is expected that this dust will 

rapidly settle within the production facility and therefore, these losses will be mainly to solid waste (EU 
PBBs and PBDEs 389 

8. POTENTIAL FOR HUMAN EXPOSURE 

2002). The compounding stage is also susceptible to dust generation, but losses are thought to be lower 

than during handling of raw materials (i.e., an order of magnitude lower). Releases will again be initially 

to the atmosphere, but the particles would be expected to settle within the compounding facility and will 

ultimately be to solid waste or waste water. In addition to particulates, emissions to the atmosphere due 

to the volatility of the flame retardant at the elevated processing temperatures may be possible. However, 

due to the high vapor pressure of decaBDE (see Table 6-6), volatilization is expected to be negligible. 

The commercial decaBDE product is an additive flame retardant used in a variety of polymer applications 

(EU 2002). The major application for decaBDE is in high-impact polystyrene (HIPS), which is used in 

the television industry for cabinet backs. Thus, decaBDE made be released during the polymer 

processing of HIPS. Waste remaining in the environment can be considered to be small particles of 

polymer product or dust generated from polymer products containing decaBDE (EU 2002). These 

particles are primarily released to soil, but may also end up in air. End products with outdoor uses are 

most likely to be atmospheric sources of decaBDE dust. Releases from these products occur over the 

lifetime of the product due to weathering and wear. Waste of this type may also be generated during 

disposal of a variety of plastic articles, in particular, where these articles are dismantled or subject to other 

mechanical processes. Ultimately, plastic articles containing technical decaBDE will be disposed of 

either to landfill or by incineration. Emissions of decaBDE from controlled incineration processes will be 

near zero (EU 2002). The concentration of decaBDE was determined in bottom ash from two municipal 

incinerators in Finland. With a method detection limit of 0.02–0.06 μg/kg, decaBDE was not detected in 

either ash sample (EU 2002). The low vapor pressure of decaBDE (see Table 6-6) limits its volatility to 

the atmosphere when disposed of in a landfill (EU 2002). 

The estimated release of 97,198 pounds (44,088 kg) of decaBDE to the atmosphere from 

54 manufacturing, processing, and waste disposal facilities in 2001 accounted for about 15.0% of the 

estimated total on-site environmental releases (TRI01 2004). These releases are summarized in 

Table 8-1. The data from the TRI listed in Table 8-1 should be used with caution, however, since only 

certain types of facilities are required to report (EPA 1995). This is not a comprehensive list. 

DecaBDE was not identified in air samples collected from 1,647 NPL hazardous waste sites (HazDat 

2004). 
PBBs and PBDEs 390 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-1. Releases to the Environment from Facilities that Produce, Process, or 

Use Decabromodiphenyl Ethera

 Reported amounts released in pounds per yearb

Statec

Number 

of 

facilities Aird Water 

Underground 

injection Land 

Total on-site 

releasee

Total offsite releasef 

Total on and 

off-site 

release 

AL 1 10 5 0 0 15 0 15 

AR 5 81,717 0 0 130,000 211,717 9,902 221,619 

CA 5 750 0 0 0 750 100,023 100,773 

CT 3 1,024 No data 0 0 1,024 24,149 25,173 

FL 2 0 No data 0 0 0 19,040 19,040 

GA 7 755 0 0 0 755 11,225 11,980 

IL 2 32 No data 0 0 32 13,700 13,732 

IN 6 20 0 0 0 20 3,987 4,007 

KY 2 555 No data 0 0 555 17,478 18,033 

LA 1 0 No data 0 270,000 270,000 0 270,000 

MA 13 292 8 0 0 300 30,658 30,958 

MD 1 0 No data 0 0 0 0 0 

MI 6 3,741 0 0 23,600 27,341 26,503 53,844 

MN 3 0 No data 0 0 0 8,507 8,507 

MO 1 500 No data 0 0 500 1,189 1,689 

MS 2 106 No data 0 0 106 30,933 31,039 

NC 14 2,818 4,237 0 99,090 106,145 73,151 179,296 

NE 2 0 No data 0 0 0 0 0 

NH 1 2 5 0 0 7 285 292 

NJ 5 97 No data 0 0 97 84,846 84,943 

NV 1 No data No data No data No data No data No data 0 

NY 4 0 5 0 990 995 36,837 37,832 

OH 9 322 0 0 750 1,072 77,277 78,349 

PA 7 1,996 109 0 0 2,105 156,439 158,544 

RI 1 13 No data 0 0 13 1,766 1,779 

SC 12 0 2,055 0 40 2,095 20,758 22,853 

TN 7 786 42 0 18,137 18,965 13,644 32,609 

TX 8 1,390 5 0 3,265 4,660 5,712 10,372 

VA 5 270 No data 0 0 270 15,087 15,357 

VT 1 0 No data 0 0 0 0 0 
PBBs and PBDEs 391 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-1. Releases to the Environment from Facilities that Produce, Process, or 

Use Decabromodiphenyl Ethera

 Reported amounts released in pounds per yearb

Statec

Number 

of 

facilities Aird Water 

Underground 

injection Land 

Total on-site 

releasee

Total offsite releasef 

Total on and 

off-site 

release 

WA 1 0 No data 0 0 0 0 0 

WI 3 2 No data 0 0 2 4,831 4,833 

Total 141 97,198 6,471 0 545,872 649,541 787,927 1,437,468 

Source: TRI01 2004 

a

The TRI data should be used with caution since only certain types of facilities are required to report. This is not an 

exhaustive list. Data are rounded to nearest whole number. 

b

Data in TRI are maximum amounts released by each facility. 

c

Post office state abbreviations are used. 

d

The sum of fugitive and stack releases are included in releases to air by a given facility. 

e

The sum of all releases of the chemical to air, land, water, and underground injection wells. f

Total amount of chemical transferred off-site, including to publicly owned treatment works (POTW). 
PBBs and PBDEs 392 

8. POTENTIAL FOR HUMAN EXPOSURE 

8.2.2 Water 

Polybrominated Biphenyls. In the past, PBBs were released to water during the manufacturing process. 

An estimated 0.0046 pounds were lost to sewers for every 1,000,000 pounds of PBBs produced at 

manufacturing sites (Neufeld et al. 1977). To manufacture PBBs, water was added to the reaction 

mixture when the desired extent of bromination was achieved. Ultimately, this water was discharged as 

effluent into surface water. Samples of effluents from the Michigan Chemical Corporation contained 

PBB concentrations 503 ppm (Di Carlo et al. 1978). Runoff water from the manufacturing plants 

containing PBBs also contaminated surface water (Di Carlo et al. 1978). Landfill sites used to dispose of 

wastes from PBB production can also be a source of PBBs in water. Concentrations of PBBs in 

groundwater from one such landfill in St. Louis, Michigan were low (0.1–0.2 ppb), but those in water 

from a drainage ditch and catch basin were much higher (0.35–1.2 ppm) (Di Carlo et al. 1978). 

PBBs have been identified in 2 and 5 surface water and groundwater samples, respectively, collected 

from 1,647 NPL hazardous waste sites (HazDat 2004). 

Polybrominated Diphenyl Ethers. Industrial and urban effluents are significant sources of PBDEs to 

surface waters and sediments. Limited data on industrial and urban effluents were located for the United 

States. Hale et al. (2002) measured the concentration of soil and stream sediments collected near a 

polyurethane manufacturing plant (near the Dan River, Virginia). Summed concentrations of 

2,2’,4,4’-tetraBDE (BDE 47), 2,2’,4,4’,5-pentaBDE (BDE 99), and 2,2’,4,4’,6-pentaBDE (BDE 100), the 

dominant congeners in these samples, ranged from <1 to 132 μg/kg (ng/g) dry weight. In 1995, sediment 

samples were collected up- and downstream near an area where the Swedish plastics industry uses 

brominated flame retardants (Sellström and Jansson 1995; Sellström et al. 1998a). Samples were 

analyzed for tetraBDEs (50 ng/g dry weight) and pentaBDEs (sum of three congeners, 2,300 ng/g dry 

weight). These PBDEs were found in higher concentrations downstream of the plant than upstream, 

which indicates that the plastics industry was the most likely source of these compounds. Surficial 

sediment samples were collected at eight locations along River Viskan near several textile manufacturing 

facilities that used various brominated flame retardants in the production of textiles. The concentrations 

of BDE 47, BDE 99, BDE 100, and BDE 209 in sediments increased as samples were collected further 

downstream where additional industries were located (Sellström et al. 1998a). The lowest levels of 

PBDEs were found upstream of the textile industries. The combined concentration of BDE 47, BDE 99, 

and BDE 100 ranged from not detected to 120 ng/g (μg/kg) dry weight; the concentration of BDE 209 

ranged from not detected to 16,000 ng/g (μg/kg) dry weight. Allchin et al. (1999) surveyed the 
PBBs and PBDEs 393 

8. POTENTIAL FOR HUMAN EXPOSURE 

concentrations of PBDEs in sediments from several rivers and estuaries in Great Britain. Sediments were 

collected upstream and downstream of suspected sources of pentaBDE and octaBDE, including a 

manufacturer, several industries, landfills, and a reference site. The highest concentrations of BDE 47, 

BDE 99, pentaBDE (as 2,3’,4’,6-tetrabromodiphenyl ether or BDE 71), and octaBDE (as 3,3’,4,5’-tetrabromodiphenyl ether or BDE 79) were in sediments near or downstream from a manufacturing site at 

Newton Aycliffe in River Skerne. The highest concentrations of decaBDE (as 2,2’,3,3’,5-pentabromodiphenyl ether or BDE 83) were found downstream of a sewage-treatment plant on River Calder. High 

concentrations were also detected on River Skerne downstream of a manufacturing site. BDE 99 

concentrations were identical or slightly higher than BDE 47 in most sediments (Allchin et al. 1999). The 

sum of five pentaBDE congeners (BDE 47, BDE 99, BDE 100, BDE 153 [i.e., 2,2’,4,4’,5,5’-hexabromodiphenyl ether], and BDE 209) ranged from 0.07 to 10.6 ng/g (μg/kg) dry weight in freshwater sediments 

from Denmark (Christensen and Platz 2001). The highest concentrations were found in sediment close to 

populated areas. 

Although the available information indicates that leaching of PBDEs from landfills is minimal, movement 

of polymer particles containing pentaBDE, octaBDE, and decaBDE commercial mixtures within the 

landfill could lead to entry into leachate water of groundwater. However, it is not currently possible to 

assess the significance of this type of process. Well designed landfills already include measures to 

minimize leaching in general, and these measures would also be effective in minimizing leaching of any 

PBDEs present (EU 2002, 2003). 

The estimated release of 6,471 pounds (2,935 kg) of decaBDE to water from nine domestic 

manufacturing and processing facilities in 2001 accounted for about 0.45% of the estimated total 

environmental releases (TRI01 2004). An additional 787,927 pounds (357,398 kg) were transferred offsite, including to POTWs (TRI01 2004). These releases are summarized in Table 8-1. The data from the 

TRI listed in Table 8-1 should be used with caution, however, since only certain types of facilities are 

required to report (EPA 1995). This is not a comprehensive list. 

PBDEs were not identified in water samples collected from 1,647 NPL hazardous waste sites (HazDat 

2004). 
PBBs and PBDEs 394 

8. POTENTIAL FOR HUMAN EXPOSURE 

8.2.3 Soil 

Polybrominated Biphenyls. The important former sources of PBBs in soil are manufacturing operations, 

disposal of PBB-containing finished products, and agricultural operations contaminated in the original 

episode in 1973–1974. The concentrations of PBBs in soils from bagging and loading areas of the 

Michigan Chemical Corporation were 3,500 and 2,500 mg/kg, respectively (Di Carlo et al. 1978). 

Similarly, soil from sites adjacent to the Hexcel Corp and the White Chemical Company, the 

manufacturers of octabromo- and decabromobiphenyl, contained decabromobiphenyl and other lower 

brominated biphenyls down to hexabromobiphenyl (Di Carlo et al. 1978). The disposal into landfills of 

solid wastes generated during the production of PBBs was another important source of PBBs in soil 

(Neufeld et al. 1977). Photodecomposition of FireMaster BP-6 in soil could also be a source of lower 

brominated biphenyls (Ruzo and Zabik 1975; Trotter 1977) in soil. 

Approximately 11.8 million pounds (5,350,000 kg) of hexabromobiphenyl was used in commercial and 

consumer products in the United States, mostly in the production of plastic products. Since the cessation 

of production of hexabromobiphenyl, all of these products, such as TV cabinet and business-machine 

housings, with a usable life of 5–10 years must have been disposed of by landfilling or incineration 

(Neufeld et al. 1977). Disposal of these plastic materials in waste-disposal sites is an important source of 

PBBs in soil. The migration of plastic-incorporated PBBs to soil would be very low since PBBs would be 

tightly bound into the plastic (Neufeld et al. 1977). 

The indirect source of PBBs in soil was the contaminated farms in Michigan. Approximately 650 pounds 

(290 kg) of PBBs was mixed in cattle feeds that were delivered to Michigan farms during 1973–1974 

(Fries 1985b). About 50% of this amount was excreted in the feces of the exposed animals and remained 

on the farms in places of fecal deposition and manure disposal (Fries 1985b). Soil in fields that received 

contaminated manure contained as high as 300 μg/kg PBBs, whereas soil in resurfaced cattle-exercise lots 

contained as high as 1,000–2,000 μg/kg of PBBs (Fries 1985b). 

PBBs have been identified in 5 soil and 3 sediment samples collected from 1,647 NPL hazardous waste 

sites (HazDat 2004). 

Polybrominated Diphenyl Ethers. PBDEs are released to land (i.e., landfills) as waste from their 

manufacture (both raw material and polymer) and as municipal wastes with the disposal of consumer 

products. Solid waste from commercial production of octaBDE is typically disposed in landfills (EPA 

1995). The disposal of consumer products containing PBDEs is likely to increase worldwide due to rapid 
PBBs and PBDEs 395 

8. POTENTIAL FOR HUMAN EXPOSURE 

obsolescence of plastic products. For example, between 1997 and 2004, the number of obsolete 

computers containing flame retardants is projected to be 315 million (NSC 1999). Based on a monitor 

weight of 30 pounds (14 kg), an estimated 350 million pounds (160,000,000 kg) of brominated flame 

retardants will be released to landfills (NSC 1999). Although PBDEs will only be a fraction of this total, 

the amount of PBDEs released to the environment by disposal will still be significant. 

PBDEs are released to farmland with their disposal as biosolids (i.e., sewage sludge). PBDEs were 

detected in biosolids destined for land applications in four different regions of the United States (Pardini 

et al. 2001). The total concentrations of pentaBDE in biosolids ranged from 1,100 to 2,290 μg/kg dry 

weight; the levels of pentaBDE were high and consistent, regardless of the region of origin. The 

concentration of decaBDE (BDE 209) varied widely among biosolids from different regions; the 

concentration of BDE 209 ranged from 84.8 to 4,890 μg/kg dry weight in the biosolid samples. 

In 2001, 545,872 pounds (247,603 kg) of decaBDE was released to land from 8 domestic manufacturing, 

processing, and waste-disposal facilities reporting releases of the compound to the environment (TRI01 

2004). No releases (0 pounds) of decaBDE occurred via underground injection (TRI01 2004). Releases 

to the environment from facilities that produce, process, or use decaBDE are summarized in Table 8-1. 

The data from the TRI should be used with caution since only certain types of facilities are required to 

report (EPA 1995). This is not a comprehensive list. 

PBDEs were not identified in soil and sediment samples collected from 1,647 NPL hazardous waste sites 

(HazDat 2004). 

8.3 ENVIRONMENTAL FATE 

8.3.1 Transport and Partitioning 

Polybrominated Biphenyls. PBBs exist predominantly in the particulate phase in the atmosphere. 

Particulate phase PBBs are removed from the atmosphere by wet and dry deposition and should not travel 

long distances in the environment. In water, PBBs are expected to absorb strongly to suspended solids 

and sediment, and may bioconcentrate in aquatic organisms. The volatilization of PBBs from water to air 

is not expected to be important due to attenuation by adsorption in the water column. In soil, PBBs are 

adsorbed strongly and will be immobile. Volatilization of PBBs from soil to air is not important due to 

the low volatility of PBBs and strong adsorption of PBBs to soil. 
PBBs and PBDEs 396 

8. POTENTIAL FOR HUMAN EXPOSURE 

Organic compounds with vapor pressures >10-4 mm Hg should exist almost entirely in the vapor phase in 

the atmosphere, while organic compounds with vapor pressures <10-8 mmHg should exist almost entirely 

in the particulate phase (Eisenreich et al. 1981). The estimated vapor pressure of FireMaster BP-6 is 

5.2x10-8 mm Hg at 25 °C (Jacobs et al. 1976). The vapor pressure of octabromobiphenyl is 7.0x10-11 mm 

Hg at 28 °C (Waritz et al. 1977). Although no data are available, the vapor pressures of 

decabromobiphenyl at ambient temperatures should be lower than octabromobiphenyl. Thus, PBBs 

produced in the 1970s should exist predominantly in the particulate phase in the atmosphere. Since the 

particulate phase PBBs would precipitate out by dry deposition and wet deposition due to washout (Atlas 

and Giam 1987), PBBs would not be expected to be transported long distances in the atmosphere. 

There are limited data regarding the transport and partitioning of PBBs in water. Based on an estimated 

Henry's law constant of 3.9x10-6 atm-m3

/mol (where Henry’s law constant = vapor pressure/water 

solubility) and an estimation method (Thomas 1990), the estimated volatilization half-life of 

hexabromobiphenyl is 23 days. Therefore, the transport of PBBs from water to the atmosphere by 

volatilization is not expected to be important. This is consistent with a fish bioconcentration study in 

which losses of octabromobiphenyl and decabromobiphenyl from water to air were found to be 

insignificant (Norris et al. 1973). Soil-mobility studies have shown that PBBs are strongly adsorbed by 

soil materials (Filonow et al. 1976; Griffin and Chou 1981a, 1981b). Therefore, sorption of water-bound 

PBBs to particulate matter and sediment is a major transport process for PBBs in water. The detection of 

at least a 1,000-fold higher concentration of PBBs in Pine River sediment (where effluent from Michigan 

Chemical Corporation was discharged) compared with the level of PBBs in the river water confirms the 

importance of this transport process (Hesse and Powers 1978). 

PBBs may also be transported from water to aquatic organisms in which bioconcentration may take place. 

Data from different laboratories on the bioconcentration of PBBs in fish show wide variation. The 

experimentally determined bioconcentration factor (BCF; the BCF is the concentration of the chemical in 

fish tissues over concentration of chemical in water) for hexabromobiphenyl (mixtures of unspecified 

congeners) in the whole body of fathead minnows (Pimephales promelas) was 18,100 in a 32-day 

exposure (Veith et al. 1979). In fillet of fathead minnow, the estimated BCF was >10,000 (Hesse and 

Powers 1978). The lipid weight-based BCF values of 4,4'-dibromobiphenyl, 2,4,6-tribromobiphenyl, 

2,2',5,5'-tetrabromobiphenyl, and 2,2',4,4',6,6'-hexabromobiphenyl in guppies (Poecilia reticulata) were 

269,000; 115,000; 1,440,000; and 708,000; respectively (Gobas et al. 1989). BCF values for mono- to 

tetra- bromobiphenyl congeners tend to increase with higher degrees of bromination while BCF values for 

tetra- and higher congeners tend to decrease with higher degrees of bromination. A similar trend in BCF 
PBBs and PBDEs 397 

8. POTENTIAL FOR HUMAN EXPOSURE 

values for various PBBs was also observed in juvenile Atlantic salmon (Salmo salar). For example, the 

whole body BCF values determined for 2,6-dibromobiphenyl, 2,4-dibromobiphenyl, 3,4-dibromobiphenyl, 2,5,4'-tribromobiphenyl, 2,2',4,5'-tetrabromobiphenyl, 2,3',4',5-tetrabromobiphenyl, 

hexabromobiphenyl (unspecified congener), and octabromobiphenyl were 1,267, 1,343, 63, 425, 314, 

111, 2–48, and 0.02, respectively (Zitko 1979; Zitko and Hutzinger 1976). The BCF values determined 

for 2,2',3,3',4,4'-hexabromobiphenyl and decabromobiphenyl in whole body guppies (P. reticulata) were 

10 and 0, respectively (Opperhuizen et al. 1985). The BCF value for octabromobiphenyl in filleted 

rainbow trout (Salmo gairdneri) was 0 (Norris et al. 1973). The lack of accumulation for the higher 

brominated compounds is most likely because they have very limited water solubility and are therefore 

not available to penetrate membranes (Zitko 1979). 

PBBs are adsorbed strongly to soil, and the adsorption increases with an increase in the organic carbon 

content of soil (Filonow et al. 1976; Griffin and Chou 1981a, 1981b). Neither clay content nor pH of soil 

correlated with adsorption of hexabromobiphenyl to soil (Filonow et al. 1976). PBBs present in soilwater solution will partition to the soil solids by adsorption. The presence of certain types of dissolved 

organic carbon in natural water (e.g., leachate from a landfill) may decrease the adsorption of PBBs in 

sediments (Simmons and Kotz 1982). Because of the strong adsorption, PBBs will have low mobility in 

soil, and the leaching of PBBs from soil to groundwater will generally be insignificant (Filonow et al. 

1976; Griffin and Chou 1981a, 1981b). However, the mobility of PBBs may greatly increase if methanol 

or other organic solvents (capable of solubilizing PBBs) are present at significant concentrations in soil as 

would happen at some contaminated sites (Griffin and Chou 1981b). This phenomenon is commonly 

called “co-solvency.” The transport of PBBs from soil to the atmosphere by volatilization is not 

important due to the low volatility and strong adsoprtion of PBBs (Jacobs et al. 1976). The transport of 

PBBs from soil to surface water or another land area via eroded soil contained in runoff water is possible 

(Jacobs et al. 1976). Orchard grass and tops of carrots grown in soil contaminated with PBBs showed no 

uptake, and carrot roots showed only minor uptake of PBBs (Jacobs et al. 1976, 1978). Therefore, the 

transport of PBBs from soil to plants via translocation is insignificant. 

Polybrominated Diphenyl Ethers. In air, PBDE commercial mixtures, which have low vapor pressures 

and exist in the particulate phase, will be removed from the atmosphere by wet and dry deposition. Thus, 

in general, PBDEs are not expected to travel long distances in the environment. However, some 

congeners in pentaBDE commercial mixtures, 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47) and 

2,2’,4,4’,5-pentabromodiphenyl ether (BDE 99), have been found in Arctic regions. It has not been 

definitively explained how these lower brominated congeners have been transported such long distances 
PBBs and PBDEs 398 

8. POTENTIAL FOR HUMAN EXPOSURE 

from areas of emission. However, it is likely that they have been transported to remote regions as dust 

particles rather than in the vapor phase. In water, PBDEs are expected to adsorb strongly to suspended 

solids and sediment, and bioconcentrate in aquatic organisms. The volatilization of PBDEs from water to 

air is not expected to be important due to attenuation by adsorption in the water column. In soil, PBDEs 

are adsorbed strongly and will be immobile. They are not likely to leach into groundwater. Volatilization 

of PBDEs from soil to air is not important due to the low volatility of PBDEs and strong adsorption of 

PBDEs to soil. 

PBDEs with vapor pressures between 10-4 and 10-8 mm Hg (di- to hexa- bromodiphenyl ether) should 

exist in both the vapor and particulate phase in the atmosphere, while PBDEs with vapor pressures 

<10-8 mm Hg (hexa- to decaBDE) should exist almost entirely in the particulate phase in the atmosphere 

(Bidleman 1988; Eisenreich et al. 1981). PBDEs have low vapor pressures, with vapor pressure tending 

to decrease with increasing bromination. Watanabe and Tatsukawa (1990) determined the vapor 

pressures for a range of brominated PBDEs as follows (mm Hg at 25 °C): di- (9.8x10-5–1.4x10-4); tri- 

(1.2x10-5–2.0x10-5); tetra- (1.8x10-6–2.5x10-6); penta- (2.2x10-7–5.5x10-7); hexa- (3.2x10-8–7.1x10-8); and 

octa- (9.0x10-10–1.7x10-9). Vapor pressures have also been determined for commercial PBDE mixtures, 

such as pentaBDE (2.2x10-7–5.5x10-7 mm Hg) and decaBDE (3.2x10-8 mm Hg at 25 °C) (EU 2001; NRC 

2000). Since particulate phase PBDEs will precipitate out by wet and dry deposition, these PBDEs would 

not be expected to be transported long distances in the atmosphere (Atlas and Giam 1987). Thus, highly 

brominated PBDEs (e.g., octa- through decaBDEs), which have low vapor pressures and exist solely in 

the particulate phase, will not be transported long distances. Although no information was located in the 

literature, moderately brominated PBDEs (e.g., pentaBDEs and hexaBDEs) may have the potential to be 

transported and would likely be found at higher concentration close to PBDE point sources. Lower BDE 

homologs (e.g., tetraBDE), which exist partially in the vapor phase, have the potential for long-range 

transport in the atmosphere (Dodder et al. 2000a). For example, 2,2’,4,4’-tetrabromodiphenyl ether 

(BDE47), 2,2’,4,4’,5-pentabromodiphenyl ether (BDE 99), and 2,2’,4,4’,5,5’-hexabromodiphenyl ether 

(BDE 153) were detected in air samples from urban, rural, and remote areas of the Great Lakes region in 

the United States (Dodder et al. 2000a). 

Concentrations of PBDEs in water are expected to be low due to the low water solubility of PBDEs. For 

example, the solubilities of the commercial mixtures of pentaBDE and decaBDE are 13.3 and <0.1 μg/L, 

respectively (EU 2001; Hardy 2002b). As consequence of their low water solubility, they have not 

detected PBDEs in environmental waters (see Section 8.4.2). 
PBBs and PBDEs 399 

8. POTENTIAL FOR HUMAN EXPOSURE 

PBDEs adsorb strongly onto suspended solids and sediments in the water column. Volatilization of 

PBDEs from water surfaces will be attenuated by adsorption, and thus is not an important fate process. 

Sediment-water partition coefficients (Kp) have been measured for several components of commercial 

pentaBDEs (Watanabe 1988). Kp values for tetra-, penta-, and hexaBDEs are 28,300, 49,200, and 

62,700 L/kg, respectively, which suggest strong partitioning to sediment. High log Kow values have been 

measured for PBDEs as follows (Watanabe and Tatsukawa 1990): di- (5.03); tri- (5.47–5.58); tetra- 

(5.87–6.16); penta- (6.46–6.97); hexa- (6.86–7.92); octa- (8.35–8.90); and deca- (9.97). Using these log 

Kow values, log organic carbon-water partition coefficients (Koc) were estimated for PBDEs: di- (4.11); 

tri- (4.35–4.41); tetra- (4.57–4.73); penta- (4.89–5.17); hexa- (5.11–5.69); octa- (5.92–6.22); and deca- 

(6.80) (Lyman et al. 1990). 

DecaBDE and octaBDE commercial products do not bioconcentrate in fish. The reported BCFs for these 

commercial mixtures are typically less than 50 (Hardy 2002b). A single study on a mixed range of 

PBDEs, between hexaBDE and decaBDE, indicated little bioconcentration in carp (e.g., Cyprinus carpio) 

with a bioconcentration factor of <4 after 8 weeks of exposure (WHO 1994a). A bioconcentration study 

was carried out with rainbow trout under static conditions. The concentration in the water was 20 μg 

14C-decaBDE per liter. Fish were exposed to decaBDE for 0, 0.5, 1, 2, 4, 6, 12, 24, or 48 hours. For each 

of the exposure periods, there was no measurable accumulation of decaBDE in flesh, skin, or viscera 

(WHO 1994a). 

An abundance of monitoring data illustrates the uptake of lower brominated diphenyl ethers by aquatic 

organisms, which results in bioconcentration (see Section 8.4.4). The commercial pentaBDE product 

undergoes bioconcentration with a BCF of approximately 14,000 (Hardy 2002b). Congener components 

of the pentaBDE commercial product bioconcentrate to different extents. For example, approximately 

50–70% of PBDEs detected in fish is a single isomer (2,2’,4,4’-tetrabromodiphenyl ether [BDE 47]). The 

next most prominent isomer is typically 2,2’,4,4’,5-pentabromodiphenyl ether (BDE 99) followed by 

2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100). In a laboratory study of Baltic blue mussels (Mytilus 

edulis L), BCFs from water absorption were found to be 1,300,000 for BDE 47, 1,400,000 for BDE 99, 

and 1,300,000 for 2,2’,4,4’,5,5’-hexabromodiphenyl ether (BDE 153) (Gustafsson et al. 1999). At several 

sites along the coast and in the Schelde estuary (the Netherlands), BCFs for blue mussels were determined 

(Booij et al. 2000). The maximum BCFs were 1x109

 for BDE 99 and BDE 100, and ≈2.5x107 for 

BDE 28, ≈2.5x108

 for BDE 47, and ≈1.6x108 for BDE 153. 
PBBs and PBDEs 400 

8. POTENTIAL FOR HUMAN EXPOSURE 

Bioaccumulation of PBDEs in the aquatic food web is inversely related to the degree of bromination 

(Burreau et al. 2000; Jansson et al. 1993). Higher brominated congeners (e.g., decaBDE) are rarely 

detected in biota. This is a result of their low solubility, high log Kow values, and sorption to soil and 

sediment (Hardy 2000). In contrast, tetraBDE to hexaBDE homologs are more frequently detected in 

biota (Burreau et al. 1997; Hale et al. 2003), which would be expected due to their greater water solubility 

and relatively high Kow values. 

Concentrations of lower brominated diphenyl ethers in biota are related to the trophic level of the species. 

For example, Haglund et al. (1997) examined the concentrations of tetra- to hexa- bromodiphenyl ethers 

in herring, salmon muscle, and gray and ringed seals collected along the Swedish coast of the Baltic Sea 

between 1981 and 1988. The concentrations of tetrabromodiphenyl ethers (e.g., 2,2’,4,4’-tetrabromodiphenyl ether, BDE 47) were found to increase with trophic level. Concentrations of PBDEs in herring 

and their predators, grey seal and guillemot, all collected at the same location of the Baltic Sea, have been 

compared to estimate potential biomagnification (de Wit 2002). The herring were caught in the autumn 

of the same year as guillemot egg collection (1987). Biomagnification factors for guillemot egg versus 

herring were 19, 17, and 7.1 for BDE 47, 2,2’4,4’,5-pentabromodiphenyl ether (BDE 99), and 

2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100), respectively. Burreau et al. (2000) analyzed small 

herring and salmon from the Atlantic Ocean (near Iceland) for several PBDEs. The calculated 

biomagnification factors for Atlantic salmon versus small herring were 3.5, 3.8, and 6.0 for BDE 47, 

BDE 99, and BDE 100, respectively. These authors concluded that biomagnification was occurring for 

the lower brominated congeners. 

PBDEs will be strongly adsorbed to soils based on log Kow values ranging from 5.03 to 9.97 (Watanabe 

and Tatsukawa 1990). Thus, PBDEs present in soil-pore water will bind to soil organic matter. Because 

PBDEs adsorb strongly to soil, they will have very low mobility (Swann et al. 1983), and leaching of 

PBDEs from soil to groundwater will be insignificant. Like PBBs, the presence of dissolved organic 

carbon in natural water may increase the mobility of PBDEs. The transport of PBDEs from soil to 

surface water via eroded soil contained in runoff water is also possible. Volatilization of PBDEs from 

moist soil surfaces will be attenuated by adsorption and is not expected to be an important fate process. 

Volatilization of PBDEs from dry soil will not be important due to the low volatility of PBDEs (see 

Table 6-6). 
PBBs and PBDEs 401 

8. POTENTIAL FOR HUMAN EXPOSURE 

8.3.2 Transformation and Degradation 

Photolysis appears to be the dominant transformation process for PBBs and PBDEs. However, the 

importance of photochemical transformation reactions in the environment cannot be determined due to 

lack of information. Based on a very limited number of studies, biodegradation does not appear to be 

significant for either PBBs or PBDEs. 

8.3.2.1 Air 

Polybrominated Biphenyls. In air, the two processes that may result in significant degradation or 

transformation of PBBs are photooxidation by hydroxyl (OH) radicals and direct photolysis. The 

estimated half-life of pentachlorobiphenyl in air due to reaction with hydroxyl radicals is 41.6–83.2 days 

(Atkinson 1987a). Based on a structure-activity relationship for the estimation of half-lives for the gasphase reactions of hydroxyl radicals with organic compounds (Atkinson 1987b), the estimated half-lives 

of hexabromobiphenyl and decabromobiphenyl due to reaction with OH radicals are 182 and 2,448 days, 

respectively. These half-lives are consistent with the half-life of pentachlorobiphenyl due to reaction with 

OH radicals. However, the half-lives of brominated biphenyls expected to be present in the particulate 

phase in the air may be even longer than the estimated half-lives due to gas phase reaction. Therefore, the 

transformation of the hexa- and other higher brominated PBBs in the atmosphere due to reaction with OH 

radicals are probably not important. 

Hexa- and other higher brominated biphenyls are expected to be present in the particle-adsorbed state in 

the atmosphere. These PBBs photolyze in solution and in soil (Hill et al. 1982; Ruzo and Zabik 1975; 

Trotter 1977). Since PBBs present in surface soil are known to photolyze, particle-sorbed PBBs present 

in the atmosphere may also undergo photolysis. The importance of the photochemical reaction under 

sunlight illumination conditions for the degradation/transformation of PBBs in air cannot be evaluated 

due the lack of information. 

Polybrominated Diphenyl Ethers. In air, PBDEs may undergo indirect photolysis with hydroxyl radicals 

or direct photolysis with sunlight. Vapor-phase PBDEs may be degraded in the atmosphere by reaction 

with photochemically produced hydroxyl radicals. The half-lives for this reaction in air are estimated to 

be 29, 140, and 476 days, respectively, for penta-, octa-, and deca- bromodiphenyl ether homologs, 

calculated using a structure estimation method (Meylan and Howard 1993). This estimation is calculated 

using an atmospheric concentration of 5x105

 hydroxyl radicals per cm3 and is based on a 24-hour day of 
PBBs and PBDEs 402 

8. POTENTIAL FOR HUMAN EXPOSURE 

sunlight. The half-lives of PBDEs that are expected to be present in the particulate phase in the air will be 

longer than the estimated half-lives calculated for the gas-phase reaction. Thus, for the higher brominated 

PBDEs (e.g., octa- and decaBDEs), indirect photolysis with hydroxyl radicals will not be important. 

In water, some PBDEs have been reported to undergo direct photolysis (Hua et al. 2003). Likewise, 

PBDEs present in the vapor phase (e.g., tetraBDE) or as particulates (e.g., decabromodiphenyl ether) may 

also undergo photolysis in the atmosphere. However, the rate and extent of the photolysis of PBDEs in 

air cannot be evaluated due the lack of information. 

8.3.2.2 Water 

Polybrominated Biphenyls. The photolytic degradation of PBBs in solution has been the subject of 

several studies. Available data in the literature indicate that brominated biphenyls photodegrade by 

reduction in solvents capable of proton transfer with the formation of lower brominated biphenyls. For 

example, the irradiation of FireMaster BP-6 and 2,2',4,4',5,5'-hexabromobiphenyl in methanol at 

wavelengths >286 nm produced mainly penta- and tetrabromobiphenyl (Ruzo and Zabik 1975). 

FireMaster BP-6 photolyzed 7 times faster than its chlorinated counterpart, 2,2',4,4',5,5'-hexachlorobiphenyl (Ruzo and Zabik 1975). Although an earlier study tentatively identified dimethoxy tetrabromobiphenyl as a photolysis product of FireMaster BP-6 (Ruzo and Zabik 1975), later work did not detect this 

compound (Ruzo et al. 1976). Earlier studies indicated that the debromination usually occurs with the 

stepwise preferential loss of bromine from the ortho and para positions of the biphenyl ring (i.e., 2, 2', 6, 

and 6' positions) (De Kok et al. 1977; Ruzo and Zabik 1975; Ruzo et al. 1976; Trotter 1977). Thus, the 

photolysis of 2,2',4,4',5,5'-hexachlorobiphenyl, the major component of FireMaster BP-6, would be 

expected to produce 2,3',4,4',5-pentabromobiphenyl and subsequently 3,3',4,4'-tetrabromobiphenyl. More 

recent work indicates that although photolysis mainly produces debromination products, unlike in the case 

of an individual PBB congener, reductive debromination of ortho substituents is not the predominant 

photolytic degradation pathway for FireMaster BP-6 (Robertson et al. 1983b). 

The study of photolysis of PBBs in the aqueous phase is more relevant to natural environmental situations 

than photolysis in proton-donating organic solvents. It was suggested that the photolysis of PBBs in 

aqueous solution would proceed by oxidative process of photohydroxylation, leading to the formation of 

phenolic compounds (Norris et al. 1973). However, photolysis of 2,4-dibromo- and 2,3',4',5-tetrabromobiphenyl in acetonitrile-water solution showed that debromination was the major reaction (Ruzo et al. 
PBBs and PBDEs 403 

8. POTENTIAL FOR HUMAN EXPOSURE 

1976). No evidence of the formation of hydroxylated species (phenolic products) was found (Ruzo et al. 

1976). 

PBBs are not expected to undergo abiotic hydrolysis under environmental conditions due to the lack of 

hydrolysable functional groups (Boethling and Mackay 2000). 

Several investigators assessed the biodegradation potential of PBBs under aerobic conditions, with 

activated sludge or pure cultures of microorganisms as microbial inoculum, and concluded that although 

the lower substituted biphenyls might biodegrade in aerobic water and sediment (Kong and Sayler 1983; 

Sugiura 1992; Yagi and Sudo 1980), the higher substituted biphenyls are resistant to aerobic 

biodegradation (Kawasaki 1980; Sasaki 1978; Shelton and Tiedje 1981). This is consistent with 

biodegradation studies in soil (see Section 8.3.2.3). It has been proposed that complete mineralization of 

4-bromobiphenyl to carbon dioxide occurs via a 4-bromobenzoate intermediate by mixed bacterial 

cultures obtained from PBB-contaminated river sediment (Kong and Sayler 1983). However, complete 

mineralization was not observed for 2- and 3-bromobenzoate (Kong and Sayler 1984). 

Although higher brominated biphenyls do not biodegrade in water or sediment under aerobic conditions, 

it has been shown that anaerobic microorganisms in river sediments obtained from populated areas can 

biodegrade higher substituted PBBs, including FireMaster mixtures (Morris et al. 1992). The 

biodegradation involved debromination at the meta and para positions, and no ortho bromine removal 

was observed (Morris et al. 1992). However, the possibility of ortho bromine removal from higher 

brominated biphenyls with certain inoculations (e.g., microorganisms from polluted river sediment 

repeatedly transferred on a pyruvate medium amended with Aroclor 1242) has been suggested (Morris et 

al. 1992). 

Polybrominated Diphenyl Ethers. PBDEs absorb light in the environmental spectrum. Hua et al. (2003) 

found that decaBDE and the commercial octaBDE absorbed light up to 325 nm, which indicates that these 

compounds may be susceptible to photodegradation at environmental wavelengths (Hua et al. 2003). Diand tetrabromodiphenyl ethers were reported to absorb minimal light at wavelengths >300 nm. This trend 

suggests that the lower brominated diphenyl ethers (e.g., pentaBDE commercial mixtures) will be less 

susceptible to photolysis compared to octaBDE and decaBDE commercial mixtures. 

PBDEs undergo debromination by direct photolysis in organic solvents and organic solvent:water 

mixtures. Laboratory studies of the photolytic breakdown of decaBDE in toluene have shown that it is 
PBBs and PBDEs 404 

8. POTENTIAL FOR HUMAN EXPOSURE 

successively debrominated by UV light to hexaBDE and that photolysis occurs very rapidly (Sellström et 

al. 1998b). The photolysis half-life in toluene was <15 minutes. However, the amounts of lower 

brominated congeners appear to be small (EU 2002). The photolysis of decaBDE (and tetra-, penta-, 

hexa-, hepta-, and octabromodiphenyl ethers) was recently reported in a 80:20 mixture of methanol:water 

at wavelengths >290 nm (EU 2002). The rate of photodegradation was found to increase with increasing 

degree of bromination. DecaBDE was found to degrade with a half-life of around 30 minutes while halflives for tetra-, penta-, hexa-, hepta-, and octabromodiphenyl ethers were 12–16 days, 2.4 days, 1.2 days, 

1.2 days, and 5 hours, respectively. The decomposition products of decaBDE were identified to be 

PBDEs (with >6 bromine atoms per molecule) and polybrominated furans (with <6 bromine atoms per 

molecule). Results of this study indicate that the photochemical stability of PBDEs increases with 

decreasing bromination (EU 2002). Rayne et al. (2003b) recently reported that 4,4’-dibromodiphenyl 

(BDE 15) photodegraded in organic (acetonitrile-methanol) and aqueous (H2O:acetonitrile; 1:1 v/v) 

solvent systems at a wavelength of 300 nm. Reductive bromination was reported to be much slower in 

the aqueous system (e.g., 73% remained after 300 minutes) compared to the organic system (where 51% 

and 41% remained after 30 minutes). However, these studies were conducted in the presence of organic 

solvents, which are not representative of conditions found in the environment. Organic solvents can act 

as hydrogen donors in photolysis reactions, which will potentially affect the distribution of products 

formed. 

In a recent study, the photolysis of PBDEs was examined under environmentally relevant conditions. 

Hua et al. (2003) studied the degradation of decaBDE in several different experiments: (1) on humic 

acid-coated silica particles, (2) on glass surfaces in contact with aqueous humic acid solutions, and (3) on 

glass surfaces in contact with water. DecaBDE dissolved in toluene was deposited on the solid substrate 

under a stream of nitrogen (to evaporate the solvent) and then desiccated to remove any residual toluene. 

The adsorbed decaBDE on the solid substrate was then inundated with the aqueous test solution, followed 

by irradiation for the duration of the test period. In all experiments, natural sunlight (location, 40° 26’ N, 

86° 55’ W) was used. The extent of degradation was determined using high-performance liquid 

chromatography (HPLC) with ultraviolet detection (UV) detection or by gas chromatography-mass 

spectrometry (GC-MS). In the first experiment, solar irradiation of decaBDE adsorbed onto humic acidsand indicated that the photolysis of decaBDE was slow. After 96 hours of exposure to sunlight, 88% of 

initial decaBDE remained on the coated sand. There is some evidence that lower brominated congeners 

(e.g., 2,2’,4,4’,6,6’-hexabromodiphenyl ether [BDE 155]) were formed in the experiment (EU 2002). In 

the second experiment, decaBDE was adsorbed on glass tubes containing a humic acid. In this study, the 

concentration of decaBDE decreased relatively quickly over the first 24 hours of exposure, after which, 
PBBs and PBDEs 405 

8. POTENTIAL FOR HUMAN EXPOSURE 

the concentration remained stable. Bromide ion accumulated at an almost linear rate from start to end of 

the 72-hour exposure period. Approximately 70% of the initial decaBDE remained after the 72-hour 

exposure. The difference in kinetics (for the disappearance of decaBDE vs. the appearance of bromide 

ion) suggests that after the initial degradation of decaBDE, bromide ion was generated by the degradation 

of lower brominated diphenyl ether congener products (possibly octa- and nonabromodiphenyl ethers). 

Bromide ion mass balance for the system indicated that 70% of the total bromine present was accounted 

for by decaBDE or bromide, with the remaining 30% present as unidentified compounds. In the third 

experiment, Hua et al. (2003) investigated the photodegradation of decaBDE adsorbed on glass tubes, 

which were filled with aqueous solutions (without humic acid). The result of this test showed a much 

more rapid loss of decaBDE than found in the analogous test using humic acid solutions. Approximately 

29% of the initial decaBDE present remained after 72 hours. The rate of decaBDE loss and bromide ion 

accumulation was relatively constant over the entire 72-hour test period. Mass balance indicated that 

approximately 50% of the total bromine was present as either decaBDE or bromide ion, while the 

remaining 50% was possibly unidentified nona- and octabromodiphenyl ether congeners. The difference 

between the tests using glass tubes with and without humic acid solution is possibly due to the absorption 

of light by humic acids, which may attenuate the degradation process. These studies indicate that 

adsorbed decaBDE may undergo photolysis forming octa- and nonabromodiphenyl ethers under 

somewhat environmentally relevant conditions. Lower brominated diphenyl ether congeners are also 

formed although only to a minor extent. These tests do not provide evidence that lower brominated 

diphenyl ethers (e.g., tetra- and pentabromodiphenyl ethers) are a major degradation product of decaBDE 

(EU 2002). There is also insufficient information from these studies to estimate the rate of photolysis or 

if intermediate degradation products build up after long-term exposures (EU 2002). 

Söderström et al. (2004) examined the time course of photolysis of decaBDE (BDE 209) in toluene, on 

silica gel, sand, sediment, and soil using artificial sunlight and on the natural matrices (e.g., sediment, 

soil, and sand) using natural sunlight. On natural samples, BDE 209 was first dissolved in toluene and 

then deposited on the natural matrix. The toluene was allowed to evaporate, and then the sample was 

reconstituted with water to resemble natural conditions. BDE 209 was photolytically labile and formed 

debromination products in all matrixes studied. Nona- to tetra- BDEs were formed as well as some 

PBDFs. The half-lives in toluene and on silica gel were less than 15 minutes, and half-lives on other 

matrices ranged from 40 to 200 hours. No differences were observed in the debromination patterns under 

different matrices or light conditions. These experiments show that photolytic debromination of BDE 209 

is a possible pathway for the formation of more bioavailable, lower brominated PBDEs. However, the 

mostt commonly found BDEs in environmental samples (e.g., 2,2’,4,4’-tetrabromodiphenyl ether 
PBBs and PBDEs 406 

8. POTENTIAL FOR HUMAN EXPOSURE 

[BDE 47], 2,2’4,4’,5-pentabromodiphenyl ether [BDE 99], and 2,2’,4,4’,6-pentabromodiphenyl ether 

[BDE 100]) were only formed to a minor degree (Söderström et al. 2004). 

Following the methodology described for decaBDE, photolysis experiments were conducted on 

2,2’,4,4’-tetraBDE (BDE 47) (EU 2002). BDE 47 was adsorbed on glass tubes filled with an aqueous 

solution and exposed to natural sunlight. After 72-hours of exposure, 30% of the initial BDE 47 

remained. The rate of disappearance of BDE 47 was comparable to that found for decaBDE under similar 

test conditions. Accumulation of bromide was initially slow with the rate increasing after 24 hours while 

the disappearance of BDE 47 was initial rapid over the first 24 hours. Using GC-MS, the authors 

concluded that 2,4,4’-tribromodiphenyl ether was being formed during this reaction and that removal of 

bromine atoms ortho to the ether functionality may be a significant reaction pathway for removal of 

bromine atoms under the conditions of this study. This study suggests that adsorbed pentabromodiphenyl 

ether congeners, like decaBDE, may undergo photolysis under somewhat environmentally relevant 

conditions (EU 2002). 

PBDEs are not expected to undergo abiotic hydrolysis under environmental conditions due to the lack of 

hydrolysable functional groups (Boethling and Mackay 2000). 

PBDEs are unlikely to biodegrade rapidly in the environment under aerobic conditions. PentaBDE did 

not undergo biodegradation (determined by CO2 evolution) after 29 days in an OECD 301B ready 

biodegradation test (EU 2001). The substance tested was a composite sample from two producers with 

the following composition: 33.7% tetraBDE, 54.6% pentaBDE, and 11.7% hexaBDE. The test was 

extended to 93 days to allow sufficient opportunity for adaptation to occur. At the end of 93 days, 2.4% 

of the theoretical amount of CO2 had been evolved. Thus, pentaBDE was determined to be not readily 

biodegradable. No degradation (as oxygen uptake) was seen for octaBDE after 28 days in an OECD 

301D ready biodegradation test (EU 2003a). Thus, octaBDE was determined to be not readily 

biodegradable. The biodegradability of decaBDE has been studied under aerobic conditions using an 

activated sludge inoculum (EU 2002). DecaBDE at 100 mg/L was incubated with activated sludge (at 

30 mg/L) over a 2-week period using a method similar to an OCED 301C MITI test. No degradation (as 

measured by biochemical oxygen demand) was observed. Thus, decaBDE was determined to be not 

readily biodegradable. 

No data on biodegradation of pentaBDE and octaBDE commercial mixtures under anaerobic conditions 

are available. An anaerobic degradation study was carried out with 2,2’,4,4’-tetrabromodiphenyl ether 
PBBs and PBDEs 407 

8. POTENTIAL FOR HUMAN EXPOSURE 

(BDE 47) using a mixture of 14C labeled and unlabeled compound (EU 2003a). The test was carried out 

using a sediment-water (Schuykill River, PA) inoculum. After 32 weeks, it appeared that no significant 

degradation of BDE 47 had occurred. However, the analytical method (i.e., HPLC using radiometric 

detection) used in this test indicated that some unidentified products had been formed in samples taken 

after 32 weeks. From these results, it is clear that BDE 47 has the potential to degrade slowly under 

anaerobic conditions (EU 2003a). Rayne et al. (2003b) recently reported that 4,4’-dibromodiphenyl 

undergoes reductive debromination under anaerobic conditions. Debromination proceeds with 

replacement of a bromine (Br) atom by a hydrogen (H) atom. The authors suggest that anaerobic 

debromination may sequentially debrominate BDE 15 to the parent diphenyl ether. 

The anaerobic biodegradability of 14C-labeled decaBDE was studied over a period of 32 weeks (EU 

2002). The test chambers consisted of 500 mL bottles containing 300 mL of sediment (Schuykill River, 

Pennsylvania) prepared under anaerobic conditions. The test chambers were incubated at 25 °C and in 

the dark during the test. After the 32-week period, <1% of the total radioactivity added was found as 

14CO2 and 14CH4 indicating that essentially no mineralization had occurred. GC-MS results showed no 

evidence for the formation of lower brominated congeners from deceBDE under the conditions of this test 

(EU 2002). 

8.3.2.3 Sediment and Soil 

Polybrominated Biphenyls. Information on the fate of PBBs in soil is limited. A pure culture of 

microorganism isolated from soil biodegraded 2-bromobiphenyl via the 2-bromobenzoic acid pathway 

(Takase et al. 1986). There is little evidence that the higher brominated biphenyls biodegrade in soil 

under aerobic conditions during an incubation period of ≤1 year (Griffin and Chou 1981a, 1981b; Jacobs 

et al. 1976). Some degradation of an undefined congener of pentabromobiphenyl was observed when 

incubated in soil, but this degradation could not be definitely attributed to biodegradation (Jacobs et al. 

1976). As discussed in Section 8.3.2.2, higher brominated PBBs may biodegrade in an anaerobic region 

of river sediment and possibly soil polluted with PCBs and PBBs to form lower brominated products. 

Biodegradation of the photolysis products of hexa- and heptabromobiphenyl in soil (which produces 

lower brominated products) was only minor (≈3% in 1 year) since the photodegradation products were 

bound to soil and light does not penetrate far into soil (Jacobs et al. 1978) 

Degradation of PBBs present in a contaminated soil from a manufacturing site in Michigan was 

significant (Hill et al. 1982). For example, 2,2',4,4',5,5'-hexabromobiphenyl, the principal component of 
PBBs and PBDEs 408 

8. POTENTIAL FOR HUMAN EXPOSURE 

FireMaster (54–68% in FireMaster) was reduced to 26% of the total PBBs when exposed to a field soil 

for several years. In two other soils, in which the original concentrations of PBBs were much lower, the 

rate of degradation was much lower. Principal degradation products were 2,3',4,4',5-pentabromobiphenyl, 

2,2',4,4',5-pentabromobiphenyl, and two unidentified tetrabromobiphenyls. The degradation was 

attributed to photochemical reactions. On the other hand, no significant photodegradation of FireMaster 

was observed after 1 year in contaminated manure spread in field soil from Michigan (Jacobs et al. 1978). 

The authors provided no explanation for the difference in photoreactivity of PBBs in soils with and 

without manure. It is important to point out that, due to attenuation and scattering of light, sunlight will 

not penetrate most soil beyond the surface layer. Therefore, it can be concluded from these studies that 

although photolysis may be the only viable degradative process for PBBs in soil, photolysis will be 

limited to the surface layer of soil, and the rate of photolysis will be very slow. PBBs incorporated into 

thermoplastics which were eventually buried at waste sites are not likely to absorb much light and 

undergo photolytic degradation. 

Analysis (Morris et al. 1993) of sediments from a PBB-contaminated river in Michigan (Pine River) 

indicates that little degradation of PBBs has occurred since the 1970s. Although microorganisms capable 

of debrominating PBBs were not present in regions of highest contamination, they were found in 

sediments downstream from the area of highest contamination. The investigators (Morris et al. 1993) 

suggest that high levels of contaminants including PBBs may be inhibiting the microbial degradation of 

PBBs in this river. 

Polybrominated Diphenyl Ethers. Information on the transformation and degradation of PBDEs in soil is 

limited. The extent to which PBDEs undergo direct photolysis in soils and sediment is unknown. 

However, sunlight would only penetrate the uppermost few millimeters of soil and will not impact 

sediment. Photolysis of PBDEs is possibily important for land-applied sewage sludge contaminated with 

PBDEs. However, no information was available on this possibility. Based on studies in water, most 

PBDEs are unlikely to biodegrade in soils or sediment under aerobic or anaerobic conditions. Lower 

brominated diphenyl ether congeners (e.g., 2,2’4,4’-tetrabromodiphenyl ether [BDE 47]) may slowly 

biodegrade under anaerobic conditions in sediment (EU 2003a). However, information on the rate or 

extent of anaerobic biodegradation of BDE 47 has not been found. No other information was found in the 

literature about the transformation and degradation processes for PBDEs in soils and sediment. 
PBBs and PBDEs 409 

8. POTENTIAL FOR HUMAN EXPOSURE 

8.3.2.4 Other Media 

No other information was found in the literature about the transformation and degradation processes for 

PBBs or PBDEs in other media. 

8.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT 

Evaluation of the potential for human exposure to PBBs or PBDEs depends in part on the reliability of 

supporting analytical data from environmental samples and biological specimens. Comparisons among 

various studies are complicated by the fact that authors may report PBB or PBDE concentrations as 

technical mixtures, as homologs, or as congeners. For PBDEs, it is common to determine the 

concentration of individual congeners. However, only a limited number of standards are available. Total 

PBDEs, the definition of what constitutes total PDBEs (i.e., how many and which congeners are 

summed), is often not the same in the various studies. Chemical analysis procedures are discussed in 

greater detail in Chapter 9. Recent monitoring data for PBBs are very limited. Historical monitoring data 

indicate that environmental PBB concentrations are confined to areas near former manufacturing facilities 

and regions of Michigan effected by the farm catastrophe of the early 1970's (see Section 8.1). 

Monitoring studies indicate that PBDEs are transported globally. Atmospheric, water, and biota levels of 

PBDEs tend to be dominated by lower brominated congeners (e.g., BDE 47). Sediments tend to be 

dominated by higher brominated congeners (e.g., BDE 209). Biota monitoring studies indicate that 

PBDE concentrations have increased since the late 1970s, with lower brominated congeners (e.g., 

BDE 47) being preferentially bioconcentrated. Studies indicate that PBDE concentrations increase with 

respect to trophic level; organisms that reside higher on the food chain tend to have higher concentrations 

of PBDEs. 

8.4.1 Air 

Polybrominated Biphenyls. Historically, PBBs were released to the atmosphere during three stages of 

the manufacturing process, and an estimate of the maximum amount of PBBs expected to be lost to the air 

during the manufacture of PBBs in the United States is available (see Section 8.2.1) (Neufeld et al. 1977). 

Monitoring data on the ambient air levels of PBBs are very limited. The concentration of 

hexabromobiphenyl in air samples collected downwind and crosswind from the White Chemical 

Company plant in Bayonne, New Jersey was 0.06 ng/m3

 (DeCarlo 1979). 
PBBs and PBDEs 410 

8. POTENTIAL FOR HUMAN EXPOSURE 

Polybrominated Diphenyl Ethers. In general, limited information is available in the literature reporting 

the levels of PBDEs in ambient air. The concentrations of PBDEs in air samples are summarized in 

Table 8-2. Atmospheric concentrations of PBDEs tend to be dominated by lower brominated congeners 

(e.g., 2,2’,4,4’-tetrabromodiphenyl ether [BDE 47], 2,2’,4,4’,5-pentabromodiphenyl ether [BDE 99], 

2,2’,4,4’,6-pentabromodiphenyl ether [BDE 100], 2,2’,4,4’,5,5’-hexabromodiphenyl ether [BDE 153], 

and 2,2’,4,4’,5,6’-hexabromodiphenyl ether [BDE 154]). DecaBDE (BDE 209) has only been detected 

in the particulate phase in air near point sources. Air samples sampled from urban (Chicago, Illinois), 

rural (Sleeping Bear Dunes, Michigan and Sturgeon Point, New York), and remote (Eagle Harbor, 

Michigan) shorelines of the U.S. Great Lakes all contained quantifiable levels of BDE 47, BDE 99, 

BDE 100, BDE 153, and BDE 154 (Dodder et al. 2000a; Strandberg et al. 2001). The most significant 

congeners were BDE 47 and BDE 99. Air measurements were averaged over a 3-year period between 

1997 and 1999. The concentration of total PBDEs ranged from 5.5 pg/m3

 in rural environments to 

52 pg/m3

 in urban air from Chicago, Illinois. The concentration of BDE 47 was 48 pg/m3 observed near 

Chicago, Illinois. The average concentration of decaBDE at the remote and rural locations was 

<0.10 pg/m3

 for each of the years investigated. The average concentration of decaBDE in the particulate 

phase at the urban location ranged from 0.20 to 0.35 pg/m3

 (Strandberg et al. 2001). 

Throughout the year of 1997, air samples were taken from a rural site in southwestern England called 

Stokes Ferry and a semirural site in northwestern England called Hazelrigg and analyzed for PBDEs 

(Peters et al. 1999). Tri- and heptabromodiphenyl ethers were detected; the combined concentrations of 

2,2’,4,4’-tetrabromodiphenyl ether (BDE 47), 2,2’,4,4’,5-pentabromodiphenyl ether (BDE 99), and 

2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100) ranged from 7 to 69 pg/m3

 at Hazelrigg and from 6 to 

58 pg/m3

 at Stoke Ferry (de Wit 2002). PBDEs have also been measured in air samples taken from 

remote stations in the Arctic (e.g., Alert, Northwest Territories, Canada; Dunai Island, eastern Siberia, 

Russia) between January 1994 and January 1995 (de Wit 2002). The total concentration of several di- to 

hexabromodiphenyl ethers ranged from 1 to 4 pg/m3

 at Alert for the majority of the year; however, in July 

1994, the concentration was 28 pg/m3

. At Dunai, the major congeners found were BDE 47 and BDE 99. 

In Sweden during 1990–1991, air samples collected from Ammarnäs in the northern mountains and 

Hoburgen on the southern tip of Gotland in the Baltic Sea, had measurable amounts of BDE 47, BDE 99, 

and BDE 100 (de Wit 2002). Total PBDE levels were approximately 1 and 8 pg/m3

, respectively. The 

concentration of BDE 47 was found to be highest in the gas phase, while BDE 99 and BDE 100 were 

highest in the particulate phase. No decaBDE was found, although the limit of detection limit for 

decaBDE is much higher than for the lower brominated diphenyl ethers. 
PBBs and PBDEs 411 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-2. Concentrations (pg/m3) of Several PBDEs in Air Samples 

Location BDE 47 BDE 99 BDE 100 ΣPBDEsa Reference 

Urban, United States 48 25 3.0 77* Dodder et al. 2000 

Rural, United States 6.2–9.2 4.3–5.0 0.6–0.9 2–4.8* Dodder et al. 2000 

Remote, United 

States 

3.7 2.6 0.33 6.9* Dodder et al. 2000 

Alert, Northwest 

Territories Canada 

No data No data No data 1–28 Alaee et al. 1999 

Eagle Harbor, 

Wisconsin 

2.9 2.1 0.28 5.5* Strandberg et al. 2001 

Sturgeon Point, New 

York 

3.8 2.8 0.39 7.2* Strandberg et al. 2001 

Sleeping Bear 

Dunes, Michigan 

8.4 5.3 0.80 15* Strandberg et al. 2001 

Chicago, Illinois 33 16 2.0 52* Strandberg et al. 2001 

Ammarnäs, Sweden 6.3 1.6 0.4 8.3 de Wit 2000, 2002 

Hoburgen, Sweden 0.7 0.35 0.07 1.1 de Wit 2000, 2002 

Stoke Ferry, United 

Kingdom 

4.7–50 5.5–13 1.1–3.9 6.7–58 Peters et al. 1999 

Hazelrigg, United 

Kingdom 

3.2–61 3.1–22 0.62–5.4 4.1–69 Peters et al. 1999 

Dunai Island, Russia No data No data No data 1–8* Alaee et al. 1999 

a

ΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an 

asterisk (*). 

Sources: de Wit 2002; Lee et al. 2002; Strandberg et al. 2001 
PBBs and PBDEs 412 

8. POTENTIAL FOR HUMAN EXPOSURE 

8.4.2 Water 

Polybrominated Biphenyls. Recent information on the concentrations of PBBs is not available. The 

concentrations of PBBs in effluents discharged from the Michigan Chemical Corporation plant in St. 

Louis, Michigan, to the Pine River during 1974–1977 ranged from <0.01 to 150 μg/L (Hesse and Powers 

1978). The concentrations of PBBs in effluents from White Chemical Company, Bayonne, New Jersey, 

and Hexcel Chemical Corporation, Sayerville, New Jersey, ranged from <0.2 to 210 μg/L (DeCarlo 

1979). The concentrations of PBBs in the Pine River ≤12 miles downstream from the Michigan Chemical 

Corporation plant in 1974 were 0.01–3.2 μg/L (Hesse and Powers 1978; Neufeld et al. 1977). 

2,2',5,5'-Tetrabromobiphenyl and 3,3',5,5'-tetrabromobiphenyl were qualitatively detected in water from 

Lake Ontario, and hexabromobiphenyl (unspecified congeners) was qualitatively detected in water from 

Lakes Ontario and Huron (Great Lakes Water Quality Board 1983). The concentrations of PBBs in test 

wells within a PBB landfill site in Michigan ranged from 0.4 to 26.0 μg/L, although the concentrations of 

PBBs in test wells outside the landfill ranged from <0.1 to 4.4 μg/L (Shah 1978). No other information 

was located about the concentrations of PBBs in water. 

Polybrominated Diphenyl Ethers. Due to the hydrophobic nature of PBDEs, this class of compounds has 

not been detected in water to any significant extent. In 1999, the concentration of PBDEs in Lake Ontario 

surface waters ranged between 4 and 13 pg/L with ~90% in the dissolved phase (Luckey et al. 2001). 

2,2’,4,4’-Tetrabromodiphenyl ether (BDE 47) and 2,2’,4,4’,5-pentabromodiphenyl ether (BDE 99) were 

the most abundant congeners, together making up >70% of the total PBDEs. In Japan, nondetectable 

levels of PBDEs were reported in 75–200 water samples (ENVIRON 2003a). No other information was 

located about the concentrations of PBDEs in water. 

8.4.3 Sediment and Soil 

Polybrominated Biphenyls. Soil samples from the bagging and loading areas of the Michigan Chemical 

Corporation plant in St. Louis, Michigan, contained PBBs at concentrations of 3,500 and 2,500 mg/kg, 

respectively (Di Carlo et al. 1978). PBBs (mostly decabromobiphenyl, but some lower brominated 

biphenyls down to hexabromobiphenyl) in soil near the Hexcel Chemical Corporation plant in New Jersey 

and the White Chemical Company plant in New Jersey ranged from 40 to 4.6 mg/kg and from 1.14 to 

4.25 mg/kg, respectively (DeCarlo 1979). PBB levels in surface soil samples from seven dairy farms in 

Michigan that spread contaminated manure on the fields ranged from 35 to 1,260 μg/kg, while the 
PBBs and PBDEs 413 

8. POTENTIAL FOR HUMAN EXPOSURE 

concentrations in surface soil of control farms (that did not use contaminated manure) were <25 μg/kg 

(Fries and Jacobs 1980). 

Concentrations of PBBs in sediments upstream from the Michigan Chemical Corporation plant were 

below the detection limit (100 μg/kg) with the exception of one sample (Hesse and Powers 1978). The 

concentration of PBBs in sediment from one upstream sample was 350 μg/kg. Hesse and Powers (1978) 

explained that this higher value was due to contamination by upstream currents during periods of waterlevel regulation at the St. Louis dam. The concentrations of PBBs in near shore sediment near the 

Michigan Chemical Corporation plant sewer outfall were ≤77.0 mg/kg. PBB concentrations in Pine River 

sediments downstream from the plant showed a gradual decrease from a maximum value of 9.2 mg/kg to 

a value of 0.1 mg/kg at a location 29 miles downstream from the plant outfall (Hesse and Powers 1978). 

Similarly, PBB concentrations in sediment samples from swamps and marshes adjacent to the White 

Chemical Company and Hexcel Chemical Corporation plants in New Jersey ranged from <10 μg/kg to 

4.6 mg/kg (DeCarlo 1979). A sludge sample from the discharge treatment plant of the White Chemical 

Company contained 431 mg/kg of PBBs (DeCarlo 1979). 

Polybrominated Diphenyl Ethers. No information was located on the ambient environmental 

concentrations of PBDEs in soils in the United States or other parts of the world. Hale et al. (2002) 

reported the concentration of PBDEs in soil samples collected in the vicinity of a polyurethane foammanufacturing facility. Levels in these soils are likely to be higher than to be expected in rural and 

potentially urban areas of the United States. Total PBDE levels in these samples ranged from not 

detected to 76 μg/kg dry weight. 2,2’,4,4’,5-Pentabromodiphenyl ether (BDE 99) was the predominant 

congener in soil followed by 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47), and 2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100). 

Sediment concentrations of PBDEs tend to be dominated by higher brominated congeners (e.g., 

BDE 209) (deWit 2002). Temporal trends suggest that concentrations of PBDEs in sediments are 

increasing. Burdens of PBDEs in sediment appear to be a function of distance from the source and their 

organic carbon content (Hale et al. 2003). The concentrations of PBDEs in sediment samples are 

summarized in Table 8-3. 

In the United States, Dodder et al. (2002) analyzed four surficial sediments from Hadley Lake (Indiana). 

This lake is in the vicinity of a production point source. DecaBDE (BDE 209) was the major congener 

detected at concentration ranging from 19 to 36 μg/kg (ng/g) dry weight. Other congeners detected (in 
PBBs and PBDEs 414 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-3. Concentrations (ng/g) of Several PBDEs in Sediment and Suspended 

Particulate Samples 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEs BDE 209 Reference 

Sediment Hadley Lake, 

Indiana 

16±2 (dw) 37±8 

(dw) 

7.1±1.5 

(dw) 

584 (dw)a 480±170 

(dw) 

Dodder et al. 

2002 

Sediment Japan No data No data No data 21–59 

(dw) 

<25–11,600 Watananbe et 

al. 1986, 1987, 

1995 

Sediment Baltic Sea ND–3.4 ND–2.4 ND–1.3 ND–5.4 ND Nylund et al. 

1992 

Sediment Upstream plastics 

plant, Sweden 

3.7 8.8 1.6 14.1 No data Sellström and 

Jansson 1995 

Sediment Downstream 

plastics plant, 

Sweden 

780 1,200 270 2,250 No data Sellström and 

Jansson 1995 

Sediment River Viskan 

(Sweden), upstream and downstream textile 

industries 

<2–50 <1–53 <0.4–19 ND–120 ND–16,000 Sellström et al. 

1998a 

Sediment 22 European river 

mouths 

<0.17–6.2 

(dw) 

<0.19–

7.0 (dw)

No data No data <0.51–1,800 de Wit 2002 

Sediment Seven rivers, 

Great Britain 

<0.3–368 

(dw) 

<0.6–

898 (dw)

No data No data <0.6–3,190 Allchin et al. 

1999 

Sediment Netherlands, 

several sites 

0.3–7.1 

(dw) 

<0.2–9 

(dw) 

No data No data <4–510 (dw) de Boer et al. 

2000b 

Suspended 

particulates 

Netherlands, 

several sites 

<2–9 (dw) <0.1–23 

(dw) 

No data No data <9–4,600 

(dw) 

de Boer et al. 

2000b 

a

includes sum of BDE 47, BDE 99, BDE 100, BDE 209, and other congeners (not specified) 

dw = dry weight; ND = not detected 

Sources: de Wit 2002; Hale et al. 2003 
PBBs and PBDEs 415 

8. POTENTIAL FOR HUMAN EXPOSURE 

decreasing order: 2,2’,4,4’,5-tetrabromodiphenyl ether [BDE 99]; 2,2’,4,4’,5,5’-hexabromodiphenyl ether 

[BDE 153]; 2,2’,4,4’,5,6’-hexabromodiphenyl ether [BDE 154]; 2,2’,4,4’-tetrabromodiphenyl ether 

[BDE 47]; and 2,2’,4,4’,6-pentabromodiphenyl ether [BDE 100]) were less than 5 μg/kg (ng/g) dry 

weight. PBDEs were above the detection limit (i.e., 0.5 μg/kg [ng/g] dry weight) in 22% of surficial 

sediment samples (from 133 sites) in freshwater tributaries of Virginia (Hale et al. 2001b). BDE 47 was 

the predominant congener followed by BDE 99 and BDE 100. The maximum concentration detected in 

sediment was 52.3 μg/kg (ng/g) dry weight. Hale et al. (2002) reported that stream sediment adjacent to a 

former polyurethane foam production facility in North Carolina contained up to 132 μg/kg (ng/g) dry 

weight of pentaBDE. 

In Japan, tetra-, penta-, hexa-, and decabromodiphenyl ethers have been observed in river sediments 

(Watanabe et al. 1986, 1987, 1995). The combined concentrations of tetra- and pentabromodiphenyl 

ethers ranged from 21–59 ng/g (μg/kg) dry weight. The concentration of decaBDE (BDE 209) ranged 

from <25 to 11,600 ng/g (μg/kg) dry weight (deWit 2002). In 1999, sediment samples from several 

locations in the Netherlands contained BDE 47, BDE 99, and BDE 209 (de Boer et al. 2000). 

Concentrations ranged from 0.3 to 7.1 ng/g (μg/kg) dry weight for BDE 47, not detected to 5.5 ng/g 

(μg/kg) dry weight for BDE 99, and not detected to 510 ng/g (μg/kg) dry weight for BDE 209. The 

concentration of PBDEs in suspended particulate matter ranged from not detected to 9 ng/g (μg/kg) dry 

weight for BDE 47, not detected to 23 ng/g (μg/kg) dry weight for BDE 99, and not detected to 

4,600 ng/g (μg/kg) dry weight for BDE 209 (de Boer et al. 2000). The concentration of several 

brominated flame retardants was measured in sediments collected from the mouths of major European 

rivers (de Wit 2002). Elevated levels of BDE 47 and BDE 99 were found in Humber and Mersey rivers 

(Great Britain). In two rivers of the Netherlands, the sum of BDE 47 and BDE 99 ranged 1.61 to 

13.1 ng/g (μg/kg) dry weight. The highest hexaBDE levels (as BDE 153) were found in the river Seine 

(France), three rivers in the Netherlands, and the rivers Schelde (Belgium), Forth (Great Britain) and Ems 

(Germany); the concentration of BDE 153 ranged from 0.013 to 0.056 ng/g (μg/kg) dry weight in these 

sediments. The concentrations of decaBDE were highest in sediment from the Seine, ranging from 2.4 to 

3.9 ng/g (μg/kg) dry weight. The concentrations of decaBDE in River Mersey (Great Britain), Schelde 

and River Liffey (Ireland) ranged from 34 to 1,800 ng/g (μg/kg) dry weight. In the southern Baltic Sea 

(Bornholm Deep), the upper layer of sediment was analyzed for BDE 47, BDE 99, and BDE 100; the 

combined concentration of these three congeners was 0.52 ng/g (μg/kg) dry weight (Nylund et al. 1992). 

A well-studied sediment core collected from the southern part of the Baltic Sea proper was analyzed for 

PBDEs and a number of organochlorine contaminants (Nylund et al. 1992). The retrospective temporal 
PBBs and PBDEs 416 

8. POTENTIAL FOR HUMAN EXPOSURE 

trend from 1939 to 1987 showed that the PBDE levels (i.e., sum of BDE 47, BDE 99, and BDE 100) have 

increased with a sharp increase after 1980. The PBDE level in the sample from 1989 was 2.9 ng/g 

(Nylund et al. 1992). Measurable amounts of BDE 28, BDE 47, BDE 66, BDE 99, and BDE 100 were 

found in sediment cores from a freshwater lake in Germany, the Wadden Sea (the Netherlands), and 

Drammenfjord (Oslo Fjord, Norway) (Zegers et al. 2000). Samples from the Drammenfjord and 

freshwater lake also contained BDE 153 and BDE 154, and the Wadden Sea and freshwater lake samples 

contained BDE 209. The lower brominated PBDEs appear in the 1960s, and BDE 209 appears about 

10 years later. The Drammenfjord sediment core shows increasing levels of BDE 47 starting in the 1940s 

(range, 0.02–0.18 ng/g dry weight) and increasing levels of BDE 99 (range, 0.5–0.28 ng/g dry weight), 

BDE 100 (range, not detected–0.07 ng/g dry weight), and BDE 154 (range, not detected–0.06 ng/g dry 

weight) beginning in the 1950s up to 1999. In the sediment core from Lake Woserin, lower brominated 

BDE congeners are detected beginning in the late 1950s, increase until the late 1970s, and then level off 

when BDE 209 first appears. A similar leveling-off trend is also observed in the Wadden Sea core 

(Zegers et al. 2000). It is important to note that this study identified the presence of PBDEs compounds 

in sediments from the late 1950s and early 1960s. This is nearly a decade prior to any significant 

commercial production of these substances. The existence of PBDEs at these early dates lends some 

credibility to the likelihood that either the substances identified in the environment as PBDEs are not 

necessarily PBDEs or that there are yet unknown sources of PBDEs produced in nature. 

8.4.4 Other Environmental Media 

Polybrominated Biphenyls. 

Food. Although the agriculture episode in Michigan involving contaminated feed occurred in May 1973, 

PBBs were not identified as the causative factor until April 1974 (Fries 1985b). PBB-containing meats, 

milk, butter, eggs, and cheese entered the human food chain for almost a year before the PBBs were 

identified. Concentrations of PBBs (on a fat basis) in milk samples collected from contaminated farms 

soon after PBB was identified ranged from 2.8 to 595 mg/kg (Cordle et al. 1978; Kay 1977). 

Concentrations of PBBs in other products processed from the contaminated milk were as follows: butter, 

1–2 mg/kg; cheese, 1.4–15.0 mg/kg; and canned milk, 1.2–1.6 mg/kg (Cordle et al. 1978). In 1974, the 

levels of PBBs in eggs from contaminated farm premises were as high as 59.7 mg/kg (Kay 1977). The 

levels of PBBs in poultry and cattle tissues from the contaminated farm collected in 1974 were 

4,600 mg/kg and up to 2,700 mg/kg, respectively (Kay 1977). With the seizure and destruction of the 

contaminated farm animals and products, the levels of PBBs in consumer products showed a steady 
PBBs and PBDEs 417 

8. POTENTIAL FOR HUMAN EXPOSURE 

decline. For example, in 1975, among 18 milk samples, 13 cheese samples, and 14 butter samples taken 

in Michigan, only 3 butter samples exceeded the FDA guidelines of 0.3 mg/kg fat (Di Carlo et al. 1978). 

In 1975, PBBs were detected in 245/2,040 meat samples collected in Michigan, with only 24 samples 

containing levels >0.3 mg/kg fat (Di Carlo et al. 1978). Although 95% of 1,430 meat samples collected 

in Michigan in 1976 contained detectable PBBs, only 1 sample contained >0.6 mg/kg, and a market 

basket survey in Michigan showed detectable PBBs in only 1/102 meat samples (Di Carlo et al. 1978). 

Fish. No PBBs were detected in several varieties of fish (carp, white sucker, Northern pike, bullhead, 

and bass) from the Alma Reservoir, which is upstream from the Michigan Chemical Corporation plant 

and above a dam that prevents fish from moving upstream (Hesse and Powers 1978). On the other hand, 

tissue samples from fish collected from the Pine River, ≤29 miles downstream from the plant, contained 

up to 1.33 mg PBBs/kg (wet weight in skinless fillets). There was no apparent change in PBB 

concentrations in fish between 1974 and 1976 (Hesse and Powers 1978). PBBs could be detected in fish 

from Pine River and other embayments and tributaries of Lake Huron in 1983. PBB concentrations in 

carp and other sedentary fish from embayments and tributaries of Lake Huron (including Pine River) and 

Lake Superior were determined (Great Lakes Water Quality Board 1989; Jaffe et al. 1985). PBBs were 

detected in the concentration range of 15–15,000 μg/kg (fat basis) in fish from embayments and 

tributaries of Lake Huron, but not from Lake Superior. Recently, Luross et al. (2002) determined the 

concentrations of several PBB congeners in lake trout from Lakes Huron, Superior, Erie, and Ontario. 

2,2’,4,4’,5,5’-Hexabromobiphenyl (BB-153) and 2,2’,4,5,5’-pentabromobiphenyl (BB-101) were found at 

the highest levels at concentrations ranging from 189 to 2,083 pg/g wet weight and from 42 to 633 pg/g 

wet weight, respectively. Several other congeners were also detected in these lake trout samples (see 

Table 8-4). 

In German rivers, elevated levels of nona- and octaBBs were present in fish. HexaBB was predominant 

in fish from the North Sea and Baltic Sea. 3,3’,4,4’,5,5’-Hexabromobiphenyl (BB-169) was found at a 

maximum concentration of 36 mg/kg (μg/g) fat in samples from the Baltic Sea. However, BB-169 was 

not found in waters from the North Sea or rivers. In Baltic marine fish, the concentrations of 

2,2’,4,4’,5,5’-hexabromobiphenyl (BB-153) ranged from 0.2 to 4.2 mg/kg (μg/g) lipid (de Boer et al. 

2000a). 

Animals. PBB concentrations in whole (with skin) and skinless ducks collected within 2 miles of the 

Michigan Chemical Corporation plant in 1974–1977 ranged from not detected to 2.70 mg/kg (μg/g) and 

not detected to 1.8 mg/kg (μg/g), respectively (Hesse and Powers 1978). Three bottlenose dolphins 
PBBs and PBDEs 418 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-4. Mean Concentrations of Nine PBB Congeners in Lake Trout from the 

Great Lakes (pg/g Wet Weight) 

Congener Lake Superior Lake Huron Lake Erie Lake Ontario 

BB-26/29 <1.3 5.2±2.3 <1.3 <1.3 

BB-31 <1.8 5.2±1.8 <1.8 <1.8 

BB-49 6.8±1.7 125±43 20±7.6 14±4.9 

BB-52 8.4±3.6 191±77 24±9.5 11±4.5 

BB-80 <3.8 <3.8 <3.8 <3.8 

BB-101 42±18 633±359 71±20 109±50 

BB-103 <1.5 4.4±1.9 <1.5 <1.5 

BB-153 189±105 2,083±1,282 220±47 1,008±513 

BB-155 1.0±0.78 5.8±3.4 <0.98 1.1±0.43 

Source: Luross et al. 2002 
PBBs and PBDEs 419 

8. POTENTIAL FOR HUMAN EXPOSURE 

(Tursiops truncatus) collected during 1987–1988 from the U.S. mid-Atlantic contained PBBs at 

concentrations of 14–20 μg/kg (ng/g) lipid basis (Kuehl et al. 1991). The source of the PBBs in the 

dolphins was not given. The median concentrations of PBBs in 10 specimens of carcass and brain of bald 

eagles (Haliaeetus leucocephalus) collected from 29 states in 1977 were 0.07 and 0.05 mg/kg (μg/g), 

respectively (Kaiser et al. 1980). Twenty-two other specimens did not contain detectable levels 

(<0.03 mg/kg [μg/g]) of PBBs. The concentrations of PBBs in eggs of fish-eating birds (common tern, 

little gull, herring gull, and red-breasted mergansers) collected during 1975–1980 from nesting islands in 

northwestern Lake Michigan and Green Bay contained PBBs in the concentration range of 0.02–

0.25 mg/kg (μg/g) wet weight (Heinz et al. 1983, 1985). 

White-tailed sea eagles collected from the Baltic Sea contained 280 ng PBBs/g lipid weight (Jansson et al. 

1987). The concentration of PBBs in common guillemots (Uria aalge) collected in 1979–1981 from the 

Baltic Sea was 160 ng/g lipid (Jansson et al. 1987). Brunnich’s guillemot (Uria lomvia), collected from 

Svalbard in the Arctic, contained 50 ng PBBs/g lipid (Jansson et al. 1987). 

In 1981, female ringed seals from Svalbard in the Swedish Arctic contained 4 ng PBBs/g lipid (Jansson et 

al. 1987). The level of PBBs in Baltic Sea harbor seal (Phoca vitulina) was 20 ng/g lipid; North Sea 

harbor seal contained 3 ng PBBs/g lipid (Jansson et al. 1987). The concentration of hexaBB ranged from 

13–61 μg/kg (ng/g) wet weight from harbor seals collected from the North Sea (decaBB <1 μg/kg [ng/g] 

wet weight). In whitebeaked dolphins from the North Sea, the concentration of hexa-, penta-, and decaBBs were 13, 8.3, and <0.9 μg/kg (ng/g) wet weight, respectively. Tetra-, penta-, and deca-BBs 

concentration ranges were 1.1–1.9, 0.4–0.9, and <0.5 μg/kg (ng/g) wet weight, respectively, in sperm 

whales from the Atlantic Ocean (de Boer et al. 1999). 

Human Tissues and Body Fluids. The quantitative determination of the concentrations of PBBs in 

blood, serum, adipose tissue, milk, and other body tissues or fluids is important in determining the human 

body burden of these chemicals. Fat is the largest repository of PBBs in the body, and concentrations in 

fat can provide an index of body burdens and exposure. It is simpler and less invasive to collect samples 

of serum or breast milk than body fat. However, the collection of milk and serum for the estimation of 

possible body burden has limitations. Breast milk can be obtained from limited segments of the 

population. Also the concentration of PBBs in breast milk can show considerable fluctuations because 

the breast is emptied only periodically (Brilliant et al. 1978; Willett et al. 1988). Serum, however, has 

lower PBB concentrations than body fat (see Section 5.5.1). 
PBBs and PBDEs 420 

8. POTENTIAL FOR HUMAN EXPOSURE 

Hexabromobiphenyl was detected (detection limit 6.6 μg/kg [ng/g]) at a frequency of 8–57% in human 

adipose tissue samples from six Canadian Great Lakes municipalities in 1984 (Williams et al. 1988). The 

concentration of 2,2',4,4',5,5'-hexabromobiphenyl in adipose tissue samples pooled from tissues of the 

general population of the conterminous Unites States ranged from 1 to 2 μg/kg (ng/g) (Lewis and 

Sovocool 1982). PBB levels in the adipose tissues of 15 quarantined dairy farm residents in midMichigan (where the mix-up involving FireMaster BP-6 occurred) ranged from 0.104 to 174 mg/kg 

(μg/g) (Humphrey and Hayner 1975). 

In the fall of 1993, the serum levels of BB-167 (2,2’,4,4’,5,5’-hexabromobiphenyl) in 32 subjects, 

approximately 10 of whom consumed sport fish from the Great Lakes, were measured (Anderson et al. 

1998). When the data were stratified by lake, on average, the Lake Huron fish consumers had the highest 

levels of PBBs (0.6 ppb [ng/g]) and Lake Erie fish consumers had the lowest (0.2 ppb [ng/g]). When the 

data were then stratified by state of residence, on average, Great Lakes sport fish consumers who live in 

Michigan had the highest PBB level (0.7 ppb [ng/g]) and residents of Wisconsin had the lowest level 

(0.05 ppb [ng/g]). 

In Michigan after the agriculture contamination episode in 1973–1974, the median PBB concentrations in 

blood of exposed adults and children in farms were 0.014 and 0.035 mg/kg (14 and 35 ppb [ng/g]), 

respectively, compared to corresponding median concentrations of 0.003 and 0.006 mg/kg (3 and 6 ppb 

[ng/g]) in a control group (Humphrey and Hayner 1975). PBB levels in the blood of quarantined farm 

workers in Michigan were also higher than in nonquarantined farm residents and the general population 

of Michigan (Cordle et al. 1978; Kimbrough 1987; Lambert et al. 1990; Landrigan et al. 1979). The 

concentration ratio of PBBs in adipose tissue over blood plasma for 13 paired specimens was 175 to 

1 (Humphrey and Hayner 1975). 

A cross-section of the population of Michigan was studied in 1978, 5 years following the agriculture 

episode involving FireMaster BP-6, to determine the levels of PBBs in human tissues. Levels of PBBs 

were highest in the part of state in which the episode occurred (median: adipose tissue, 500 μg/kg (ng/g); 

serum, 1.7 μg/L) and were lowest in the upper peninsula (median: adipose tissue, 15 μg/kg (ng/g); serum, 

0.2 μg/L), farthest from the source of contamination. Levels in the rest of the state were in between 

(median: adipose tissue, 240 μg/kg [ng/g]; serum, 0.9 μg/L) (Wolff et al. 1982). The estimated 

concentration ratio of PBBs in adipose tissue over serum was near 300 among 31 Michigan dairy farm 

residents (Wolff et al. 1979a). The ratio of adipose tissue to serum PBB concentration was 363 to 1 for 

the general population and 100 to 1 in lactating women (Brilliant et al. 1978). The kinetics of fat 
PBBs and PBDEs 421 

8. POTENTIAL FOR HUMAN EXPOSURE 

metabolism in lactating women seems to alter PBB partitioning. The ratios of adipose tissue to serum 

PBB concentration for nonpregnant females and male chemical workers, farm workers and other males, 

and pregnant females in 3,683 Michigan residents with varying degrees of exposure were 190–260 to 1, 

325–329 to 1, and 107–119 to 1, respectively (Eyster et al. 1983). The PBB ratios for cord to serum and 

placenta to serum in pregnant females were 0.10–0.14 to 1 and 0.10–0.17 to 1, respectively (Eyster et al. 

1983). The PBB ratios for feces to serum and bile to serum in farm and chemical workers were 0.53–

0.71 to 1 and 0.45–0.63 to 1, respectively (Eyster et al. 1983). The detection of PBBs in bile and feces 

indicates transfer into the intestinal tract. However, the concentration of PBBs in feces represented a 

minor proportion of the total body burden, indicating a slow rate of excretion (Eyster et al. 1983). Serum 

PBBs were determined in a nested case-control study of 1,925 women enrolled in the Michigan 

Department of Public Health registry for persons exposed to PBBs (Henderson et al. 1995). Study 

participants had lived on or received food from a farm quarantined by the Michigan Department of 

Agriculture, were recruited from July 1976 to December 1977, and were followed up annually from 1978 

through 1993. Median serum PBB concentrations were 2 ppb (n=290; range=0.5–419 ppb [μg/L]). 

The concentrations of PBBs in the breast milk of females from the lower peninsula of Michigan (exposed 

area) were generally higher than in breast milk of females from the upper peninsula (farthest from the 

sources) (Brilliant et al. 1978). PBB levels in breast milk of five females from the exposed farms were 

0.21–92.7 mg/kg (Cordle et al. 1978; Humphrey and Hayner 1976). In a cohort of Michigan residents, 

the ratio of PBBs in breast milk to maternal serum was 107–122 to 1 and in adipose tissue to breast milk 

was 1.1–1.5 to 1 (Eyster et al. 1983; Landrigan et al. 1979). The concentrations of PBBs found in human 

tissues and body fluids are given in Table 8-5. Recent levels of PBBs in human breast milk (i.e., 1990 to 

present) were not located (WHO 1994b). 

Cow’s Milk. In an attempt to determine the metabolites of PBBs, whole milk of lactating cows from 

contaminated areas of Michigan was analyzed for monohydroxy metabolites, but none were found 

(Gardner et al. 1976). In a later study, the feces of dogs fed FireMaster BP-6 in corn oil was found to 

contain a metabolite identified as 6-hydroxy-2,2',4,4',5,5'-hexabromobiphenyl (Gardner et al. 1979). 

The effects of processing cow's milk containing PBBs also has been studied (Murata et al. 1977; Zabik et 

al. 1978). Spray-drying reduced PBB levels in whole and skim milk, whereas pasteurization, freezedrying, aging of cheese, and condensation were not effective in reducing the level of PBBs in milk 

products. Pressure-cooking meat containing PBBs reduced the level of PBBs in the cooked meat (Zabik 

et al. 1978). 
PBBs and PBDEs 422 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-5. Tissue Levels of PBBs in Michigan Residents 

Tissue Subject(s) 

Mean/median 

concentrationa Year Reference 

Serum Exposed farm workers 14 μg/L 1976 Stross et al. 1979 

 Chemical workers 48 μg/L No data Stross et al. 1981 

 Chemical workers 1.1–1,000 μg/L 1976 Anderson et al. 1978d

 Exposed farm workers BDL to 

1,000 μg/L 

1976 Anderson et al. 1978d

 Residents from quarantined farms 26.9 μg/L 1976–1977 Landrigan et al. 1979 

 Residents from non-quarantined 

farms 

3.5 μg/L 1976–1977 Landrigan et al. 1979 

 Farm product consumers 17.1 μg/L 1976–1977 Landrigan et al. 1979 

 Chemical workers and families 43.0 μg/L 1976–1977 Landrigan et al. 1979 

 Control group 3.5 μg/L 1976–1977 Landrigan et al. 1979 

 General population (lower peninsula) 1.9 μg/L 1978 Wolff et al. 1982 

 General population (upper 

peninsula) 

0.2 μg/L 1978 Wolff et al. 1982 

 General population (remainder of 

state) 

0.9 μg/L 1978 Wolff et al. 1982 

 Chemical workers 25.4 μg/L No data Eyster et al. 1983 

 Farm and other workers 5.4 μg/L No data Eyster et al. 1983 

 Mothers from lower peninsula 26.2 μg/L 1976–1977 Landrigan et al. 1979 

 Exposed mothers from farms 3.4 μg/L No data Eyster et al. 1983 

 Non-pregnant women from exposed 

farms 

3.1 μg/L No data Eyster et al. 1983 

 Exposed women enrolled in the 

Michigan Department of Public 

Health registry 

2 μg/L 1993 Henderson et al. 1995

Cord serum Exposed mothers from lower 

peninsula 

3.2 μg/L 1976–1977 Landrigan et al. 1979 

 Mothers from lower peninsula <1.0 μg/L No data Eyster et al. 1983 

Blood plasma Workers from quarantined farms 14 μg/L 1974 Humphrey and 

Hayner 1975 

 Children from quarantined farms 35 μg/L 1974 Humphrey and 

Hayner 1975 

 Adults from non-quarantined farms 3 μg/L 1974 Humphrey and 

Hayner 1975 

 Children from non-quarantined farms 6 μg/L 1974 Humphrey and 

Hayner 1975 

Placenta Exposed mothers <1 μg/L No data Eyster et al. 1983 

Breast milk Exposed mothers 370 μg/kg (fat 

basis) 

No data Eyster et al. 1983 

 Exposed mothers from lower 

peninsula 

3,614 μg/kg (fat 

basis) 

1976–1977 Landrigan et al. 1979 

 Mothers from lower peninsula 68 μg/kg (fat 

basis) 

1976 Brilliant et al. 1978 

 Mothers from upper peninsula <44 μg/kg (fat 

basis) 

1976 Brilliant et al. 1978 
PBBs and PBDEs 423 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-5. Tissue Levels of PBBs in Michigan Residents 

Tissue Subject(s) 

Mean/median 

concentrationa Year Reference 

Adipose 

tissue 

Population of lower peninsula 500 μg/kg 1978 Wolff et al. 1982 

 Population of upper peninsula 15 μg/kg 1978 Wolff et al. 1982 

 Population of rest of the state 240 μg/kg 1978 Wolff et al. 1982 

 Chemical workers 9,330 μg/kg No data Brown et al. 1981 

 Farm residents from lower peninsula 3,940 μg/kg No data Brown et al. 1981 

 Farm residents from lower peninsula 3,260 μg/kg 1976 Stross et al. 1979 

 Chemical workers 12,820 μg/kg No data Stross et al. 1981 

 Workers from quarantined dairy 

farms 

12,500 μg/kg 1974 Humphrey and 

Hayner 1975 

 Pregnant females from lower 

peninsula 

400 μg/kg No data Eyster et al. 1983 

 Chemical workers 5,290 μg/kg No data Eyster et al. 1983 

 Farm and other workers from lower 

peninsula 

1,650 μg/kg No data Eyster et al. 1983 

a

When both mean and median values are available, the former values have been used in the table. In some cases, 

when neither value is available, the range is given in the table. 

BDL = below detection limit 
PBBs and PBDEs 424 

8. POTENTIAL FOR HUMAN EXPOSURE 

Polybrominated Diphenyl Ethers. 

Food. Information about the concentrations of PBDEs in food-stuff is very limited. Huwe et al. (2002a) 

reported total PBDE levels in farm chickens raised in two different regions of the United States. The total 

PBDE level of discrete samples of chickens raised in Arkansas was 39.4 ng/g whole weight, while one 

composite sample of chickens raised in North Dakota was 1.7 ng/g whole weight. In the United States, 

chickens fed ball clay and chickens bought in the grocery store were analyzed for total PBDEs (Environ 

2003b). 2,2’,4,4’,5-Pentabromodiphenyl ether (BDE 99) was the dominant congener in all samples. 

Total PBDEs ranged between 4 and 35 ng/g lipid weight in chickens fed ball clay and 0.5 ng/g lipid 

weight in store-bought chicken. Recently, Ohta et al. (2002a) determined the concentration of total 

PBDEs in vegetables and meat samples from Japan. The concentrations of PBDEs in spinach, potato, and 

carrot were 134, 47.6, and 38.4 pg/g fresh weight, respectively. The highest concentrations of total 

PBDEs and 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47) were found in spinach. Interestingly, different 

congener patterns were found among the vegetables analyzed. Compared to root vegetables, which had 

high concentrations of 2,2’,4,4’,5,5’-hexabromodiphenyl ether (BDE 153), spinach (representing a leafy 

vegetable) might be strongly influenced by PBDE contamination in air. The concentration of PBDEs in 

pork, beef, and chicken were 63.6, 16.2, and 6.25 pg/g fresh weight, respectively. PBDE concentrations 

were highest in pork samples; however, the reason for this is unknown (Ohta et al. 2002a). Bocio et al. 

(2003) determined the concentrations of PBDEs in food samples from Catalonia, Spain during 2000. The 

highest concentration of total PBDEs was found in oils and fats (587.7–569.3 pg/g), followed by fish and 

shellfish (333.9–325.3 pg/g), meat and meat products (109.2–102.4 pg/g), and eggs (64.5–58.3 pg/g). In 

all these food groups, a predominance of the tetra- and pentaBDE homologs, followed by hexaBDE, was 

observed in the sum total PBDEs. By contrast, PBDEs were not detected in the groups of fruits, cereals, 

and tubers. Four types of commercial fish oils sold in Sweden were found to contain PBDEs (0.2–

28.1 ng/g lipid weight) (Haglund et al. 1997). The highest concentration of PBDEs was found in the cod 

liver oil. These oils were from products marketed as dietary supplements for humans. The concentrations 

of PBDEs in seafoods from the Inland Sea of Japan were determined for samples collected in 1998 (Hori 

et al. 2000). The congeners, 2,4,4’-tribromodiphenyl ether (BDE 28); BDE 47; 2,3’,4,4’-tetrabromodiphenyl ether (BDE 66); BDE 99; 2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100); BDE 153; and 

2,2’,4,4’,5,6’-hexabromodiphenyl ether (BDE 154) were detected in all analyzed seafood samples. 

BDE 47 was detected as the predominant congener with concentrations ranging from 58 to 2,100 pg/g wet 

weight. Harrad et al. (2004) determined the concentrations of several PBDE concengers in omnivorous 

and vegetarian diet samples from the United Kingdom. Median concentrations of BDE 47 
PBBs and PBDEs 425 

8. POTENTIAL FOR HUMAN EXPOSURE 

(2,2’,4,4’-tetraBDE), 99 (2,2’,4,4’,5-pentaBDE), 100 (2,2’,4,4’,6-pentaBDE), 153 (2,2’,4,4’,5,5’-hexaBDE), and 154 (2,2’,4,4’,5,6’-hexaBDE) in omnivorous diet samples were 66.8, 63.8, 10, 20, and 20 pg/g 

dry weight, respectively. In vegetarian samples, median concentrations of BDE 47 (2,2’,4,4’-tetraBDE), 

99 (2,2’,4,4’,5-pentaBDE), 100 (2,2’,4,4’,6-pentaBDE), 153 (2,2’,4,4’,5,5’-hexaBDE), and 

154 (2,2’,4,4’,5,6’-hexaBDE) were 47.2, 56.7, 10.0, 20, and 20 pg/g dry weight, respectively. 

Concentrations of BDE 47, 99, and total PBDE were found to be statistically higher in omnivorous diet 

samples compared to vegitarian diet samples. 

Biosolids and Effluents. The concentrations of PBDEs in biosolids (sewage sludge) and effluents are 

summarized in Table 8-6. PBDEs were detected in biosolids from four different regions of the United 

States (Pardini et al. 2001). The total concentrations of pentaBDE in biosolids ranged from 1,100 to 

2,290 ng/g dry weight; the levels of pantaBDE were high and consistent, regardless of the region of 

origin. The concentration of decaBDE (BDE 209) varied widely among biosolids from different regions; 

the concentration of BDE 209 ranged from 84.8 to 4,890 ng/g dry weight in the biosolid samples. 

Sewage sludges in the vicinity of the Dan River (Virginia) were collected and analyzed for PBDEs (Hale 

et al. 2002). Congener patterns suggestive of both penta- and decaBDE commercial products were 

present at concentrations of 1,370 ng/g dry weight (sum of BDE 47 to BDE 154) and from 1,470 ng/g dry 

weight, respectively. While no known industrial source of pentaBDE discharged to this plant, the 

distribution pattern for lower brominated congeners matched the pentaBDE commercial product. 

Sewage sludge samples from 13 waste water-treatment plants in Germany were sampled (Hagenmaier et 

al. 1992). The mean concentration of tri- to hepta-BDEs was 8.37 ng/g with tri-, tetra-, penta-, hexa-, and 

heptaBDE at concentrations of 0.65, 3.06, 3.02, 0.49, and 0.22 ng/g, respectively. Levels of penta- and 

hexaBDEs were highest in these samples. de Boer et al. (2000b) determined the concentration of PBDEs 

in sewage treatment-plant effluents (STP) from the Netherlands. The concentration of total PBDEs 

ranged from 11 to 35 ng/g dry weight, while the concentrations of 2,2’,4,4’-tetrabromodiphenyl ether 

(BDE 47) and BDE 209 were 11–35 and 310–920 ng/g dry weight, respectively. Kohler et al. (2003) 

determined the levels of decaBDE in sewage sludge from Switzerland between 1993 and 2002. These 

authors reported that the average concentration of decaBDE increased with time from 220 to 1,100 ng/g 

dry weight, corresponding to an average increase of 560%. 

World Trade Center Site. In 2001, PBDEs were detected in dust and smoke samples taken near the 

World Trade Center (WTC) disaster site (Lioy et al. 2002). The highest concentration was for decaBDE 

(i.e., BDE 209), which was present in thermoplastics (e.g., computers). Concentrations of PBDE 
PBBs and PBDEs 426 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-6. Concentrations (ng/g Dry Weight) of Several PBDEs in Biosolids 

(Sewage Sludge) and Effluents 

Sample type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 Reference 

Sewage 

sludge 

Dan River, 

Virginia 

No data No data No data 2,840* 1,470 Hale et al. 2002 

Sewage 

sludge 

Gothenburg, 

Sweden 

15 19 3.5 38 No data Nylund et al. 

1992 

Sewage 

sludge 

Klippan, 

Sweden 

22 18 5.4 45.4 No data Sellström 1999; 

Sellström and 

Jansson 1995 

Sewage 

Sludge 

Rimbo, 

Sweden 

53 53 13 119 No data Sellström 1999; 

Sellström and 

Jansson 1995 

Sewage 

sludge 

Three plants, 

Stockholm, 

Sweden 

39–91 48–120 11–28 98–239 140–350 Sellström et al. 

1999 

Sewage 

sludge 

Germany No data No data No data 04–15* No data Hagenmaier et 

al. 1992 

Sewage 

treatment 

plant effluents 

Netherlands, 

several sites 

11–35 <1 No data 11–35 310–920 de Boer et al. 

2000b 

a

ΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an 

asterisk (*). 

Source: de Wit 2002; Hale et al. 2002 
PBBs and PBDEs 427 

8. POTENTIAL FOR HUMAN EXPOSURE 

congeners were 107–174 μg/kg dry weight basis for 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47), 51.1–

74.1 μg/kg dry weight basis for 2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100), 155–293 μg/kg dry 

weight basis for 2,2’,4,4’,5-pentabromodiphenyl ether (BDE 99), 42.0–53.5 μg/kg dry weight basis for 

2,2’,4,4’,5,5’-hexabromodiphenyl ether (BDE 153), 219–305 μg/kg dry wt basis for 2,2’,4,4’,5,6’-hexabromodiphenyl ether (BDE 154)/PBB-153, and 1,330–2,660 μg/kg dry wt basis for decaBDE (BDE 209). 

Levels of PBDEs were found to be similar to levels found in sewage sludge (Lioy et al. 2002). No further 

information on levels of PBDEs in environmental was located for the WTC site. 

Freshwater Fish. Monitoring data indicated that the levels of PBDEs are increasing in freshwater 

organisms with higher concentrations near point sources. The congener profiles show the highest levels 

for 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47). The presence of PBDEs in freshwater aquatic 

organisms taken from remote regions suggests that diffuse sources of PBDEs are also important. The 

concentrations of PBDEs in freshwater fish samples in the United States are summarized in Table 8-7. 

Fish were sampled from two U.S. lakes, Hadley Lake, Indiana near a possible PBDE point source, and 

Lake of the Ozarks, Missouri, with no known sources (Dodder et al. 2000a). Mean total PBDE 

concentrations (sum of BDE 47, 2,2’,4,4’,5-pentabromodiphenyl ether [BDE 99], 2,2’,4,4’,6-pentabromodiphenyl ether [BDE 100], 2,2’,4,4’,5,5’-hexabromodiphenyl ether [BDE 153], and 2,2’,4,4’,5,6’-hexabromodiphenyl ether [BDE 154]) were higher in crappie (Poxomis annularis) and bluegill (Lepomis 

macrochiras) from Hadley Lake (1,500 and 1,900 ng/g lipid weight, respectively) than from Lake of the 

Ozarks (340 and 390 ng/g lipid, respectively). BDE 47, BDE 99, BDE 153, and BDE 154 were primary 

congeners. From the Lake of Ozarks, BDE 47 was the dominant congener in fish. The total PBDE 

concentrations in smelt (Osmerus mordax) from Lakes Superior and Ontario were 150±9 and 

240±30 ng/g lipid, respectively (Dodder et al. 2002). The dominate congeners in these fish were BDE 47 

and BDE 99. An analysis of fish tissue samples from selected locations in Washington State showed that 

total PBDE concentrations ranged from 29 ng/g lipid in rainbow trout from a remote spring-fed stream 

(Douglas Creek, Washington) to 19,000 ng/g lipid in rainbow trout from the urbanized Spokane River, 

Washington (Johnson and Olson 2001). The tetra- and pentaBDE isomers were the major compounds 

present. TetraBDE to hexaBDE were found in carp (Cyprinus carpio) from the Buffalo River (New 

York), a polluted area around the Great Lakes (Loganathan et al. 1995). TetraBDEs dominated the 

congener pattern with 94–96% of total PBDEs. TetraBDE and pentaBDE concentrations ranged from 

13 to 22 ng/g fresh weight. Asplund et al. (1999a) found tri- to hexaBDEs in steelhead trout 

(Oncorhynchus mykiss) sampled in 1995 from Lake Michigan. The combined concentration of BDE 47, 

BDE 99, BDE 100, BDE 153, and BDE 154 was 3,000 ng/g lipid weight (Asplund et al. 1999b). Lake 

trout (Salvelinus manaycush) from Lakes Ontario, Huron, and Superior were also found to have di- to 
PBBs and PBDEs 428 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-7. Concentrations (ng/g Lipid Weight, Except as Noted) of Several PBDEs 

in Freshwater Fish Samples from the United States 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa

 BDE 209 Reference 

Alewife Grand Traverse 

Bay, Lake 

Michigan 

16 No data No data 36 No data Stapleton and 

Baker 2003 

Bloater 

chub 

Grand Traverse 

Bay, Lake 

Michigan 

11 (fw) No data No data 23 (fw) No data Stapleton and 

Baker 2003 

Bluegill Hadley Lake, 

Indiana 

420 320 240 1,900 No data Dodder et al. 

2000 

Bluegill Lake of the 

Ozarks, Missouri 

200 91 59 390 No data Dodder et al. 

2000 

Burbot Grand Traverse 

Bay, Lake 

Michigan 

43 (fw) No data No data 86 (fw) No data Stapleton and 

Baker 2003 

Carp United States No data No data No data 13–22* (fw) No data Loganathan et 

al. 1995 

Carp Detroit River, 

Grosse Isle, 

Michigan 

3.0 (fw) 0.50 (fw) 0.48 (fw) 40.7* No data Rice et al. 2002 

Carp Des Plaines 

River, Joliet, 

Illinois 

2.54 (fw) 0.5 (fw) 0.44 (fw) 281* No data Rice et al. 2002 

Carp Des Plaines 

River, Joliet, 

Illinois 

1.34 (fw) 0.50 (fw) 0.49 (fw) 78.3* No data Rice et al. 2002 

Carp (fillet) Yakima River, 

Washington 

No data No data No data 22 (fw) No data Johnson and 

Olson 2001 

Crappie Hadley Lake, 

Indiana 

250 430 150 1,500* No data Dodder et al. 

2000 

Crappie Lake of the 

Ozarks, Missouri 

190 78 59 340* No data Dodder et al. 

2000 

Deepwater 

sculpin 

Grand Traverse 

Bay, Lake 

Michigan 

2.8 (fw) No data No data 3 (fw) No data Stapleton and 

Baker 2003 

Lake trout Grand Traverse 

Bay, Lake 

Michigan 

75 (fw) No data No data 126 (fw) No data Stapleton and 

Baker 2003 

Lake trout Lake Ontario, 

United States 

No data No data No data 540* No data Alaee et al. 1999

Lake trout Lake Ontario, 

United States 

58 (fw) 14 (fw) 5.7 (fw) No data No data Luross et al. 

2002 

Lake trout Lake Huron, 

United States 

No data No data No data 240* No data Alaee et al. 1999

Lake trout Lake Huron, 

United States 

27 (fw) 7.7 (fw) 3.8 (fw) No data No data Luross et al. 

2002 

Lake trout Lake Superior, 

United States 

No data No data No data 140* No data Alaee et al. 1999

Lake trout Lake Superior, 

United States 

29 (fw) 12 (fw) 4.1 (fw) No data No data Luross et al. 

2002 

Lake trout Lake Erie, 

United States 

No data No data No data 117* No data Alaee et al. 1999
PBBs and PBDEs 429 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-7. Concentrations (ng/g Lipid Weight, Except as Noted) of Several PBDEs 

in Freshwater Fish Samples from the United States 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa

 BDE 209 Reference 

Lake trout Lake Erie, 

United States 

16 (fw) 2.0 (fw) 2.5 (fw) No data No data Luross et al. 

2002 

Largescale 

sucker 

(whole) 

Yakima River, 

Washington 

No data No data No data 64 (fw) No data Johnson and 

Olson 2001 

Largescale 

sucker 

(whole) 

Spokane River, 

Washington 

No data No data No data 105 (fw) No data Johnson and 

Olson 2001 

Mountain 

whitefish 

(whole) 

Spokane River, 

Washington 

No data No data No data 1,250 (fw) No data Johnson and 

Olson 2001 

Rainbow 

trout 

(whole) 

Douglas Creek, 

Washington 

No data No data No data 1.5 (fw) No data Johnson and 

Olson 2001 

Rainbow 

trout 

Spokane River, 

Washington 

No data No data No data 20-174 (fw) 

(fillet) 

297 (fw) 

(whole) 

No data Johnson and 

Olson 2001 

Salmon Grand Traverse 

Bay, Lake 

Michigan 

34 (fw) No data No data 95 (fw) No data Stapleton and 

Baker 2003 

Salmon Lake Michigan, 

United States 

52.1 (fw) 9.3 (fw) 9.7 (fw) 2,440 No data ManchesterNeesvig et al. 

2001 

Smelt Lake Superior, 

United States 

5.7 (fw) 1.8 (fw) 0.98 (fw) 150 <1.5 (fw) Dodder et al. 

2002 

Smelt Lake Ontario, 

United States 

10 (fw) 5.3 (fw) 1.6 (fw) 240 <1.6 (fw) Dodder et al. 

2002 

Starry 

flounder 

(whole) 

Columbia River, 

Washington 

No data No data No data 30 (fw) No data Johnson and 

Olson 2001 

Steelhead 

trout 

Lake Michigan, 

United States 

1,700 600 360 3,000* No data Asplund et al. 

1999b 

Whitefish Columbia River, 

United States 

No data No data No data 72 (fw) No data Rayne et al. 

2003a 

Whitefish Grand Traverse 

Bay, Lake 

Michigan 

9.8 (fw) No data No data 18 (fw) No data Stapleton and 

Baker 2003 

a

ΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an 

asterisk (*). 

Sources: Luross et al. 2002; Manchester-Neesvig et al. 2001; Rayne et al. 2003; Stapleton and Baker 2003 

dw = dry weight; fw = fresh weight 
PBBs and PBDEs 430 

8. POTENTIAL FOR HUMAN EXPOSURE 

heptaBDEs with combined concentrations of 545, 237, and 135 ng/g lipid weight, respectively (Alaee et 

al. 1999). Lake trout from Lake Erie had 117 ng/g lipid weight (Luross et al. 2000). Variations in local 

sources, combined with atmospheric transport, may explain differences that were seen in congener 

profiles for the different lakes. A retrospective temporal study for the years 1978, 1983, 1988, 1993, and 

1998 using archived trout samples from Lake Ontario show a dramatic increase in total PBDE 

concentrations over time (Luross et al. 2000). At 50 fresh water sites in Virginia, muscle samples from 

253 fish samples were collected and analyzed for PBDEs (Hale et al. 2000, 2001b). Approximately 85% 

of the samples contained BDE 47, the predominant congener, at measurable concentrations. 

Concentrations were >1,000 ng/g lipid weight at 9 of 50 sites. The highest combined PBDE 

concentrations (up to 57,000 ng/g lipid weight) were observed in carp downstream of textile and furniture 

facilities. BDE 47 levels were greater than 2,2’,4,4’,5,5’-hexachlorobiphenyl (PCB 153) concentrations 

in 58% of the samples analyzed. PBDEs were identified in fish collected from the Detroit River 

(Michigan) and Des Plaines Rivers (Illinois). In Detroit River fish (carp and large mouth bass), the 

congener patterns were dominated by BDE 47; however, in the Des Plaines River carp, the dominant 

congeners were heptaBDE congeners (2,2’,3,4,4’,5,6-heptBDE [BDE 181] and 2,2’,3,4,4’,5’,6-heptaBDE 

[BDE 183]), lesser amounts of 2,3,3’,4,4’,5,6-heptaBDE (BDE 190), and two hexaBDEs (BDE 154 and 

BDE 153). Possible sources for the heptaBDE congeners were not obvious since none of the commercial 

mixtures are known to contain these congeners. Three possible explanations were proposed to explain the 

presence of the heptaBDE congeners found in Des Plaines River carp: waste discharge from an industrial 

facility; publicly-owned treatment works (POTW) effluents; or formation in situ by decaBDE deposits. 

The concentrations of PBDEs in freshwater fish samples from Europe are summarized in Table 8-8. 

Between 1986 and 1988, levels of 2,2’4,4’-tetrabromodiphenyl ether (BDE 47), 2,2’,4,4’,5-pentabromodiphenyl ether (BDE 99), and 2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100) were measured in 

whitefish (Coregonus spp.) from a remote mountain lake in Northern Sweden (Lake Storvindeln), Arctic 

char (Salvelinus alpinus) from a heavily populated lake (Lake Vättern) in south-central Sweden with 

numerous municipal and industrial point sources, and in trout (Salmo trutta) and pike (Esox lucius) from 

several sites along Dalslands Canal in west central Sweden (Jansson et al. 1993). No point sources of 

PBDEs were identified from these sites. Whitefish from the remote lake contained the lowest levels 

(26 ng/g lipid weight) of PBDEs, whereas the Arctic char, from a heavily populated lake, contained 

520 ng/g lipid weight PBDEs. In both samples, BDE 47 was the predominant congener. PBDE 

concentration levels in pike and trout from the Dalslands Canal ranged from 180 to 210 ng/g lipid weight 

and from 280 to 1,200 ng/g lipid weight, respectively. The congener pattern in these samples was similar 

to the technical mixture, Bromkal 70-5DE, with equal quantities of both BDE 47 and BDE 99. The levels 
PBBs and PBDEs 431 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-8. Concentrations (ng/g Lipid Weight Except as Noted) of Several PBDEs 

in Freshwater Fish Samples from Europe 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa

 BDE 209 Reference 

Arctic char Lake Vättern, 

Sweden 

400 64 51 520 No data Sellström et al. 

1993 

Bream Netherlands 

(several sites) 

0.2–130 

(dw) 

Not 

detected 

No data No data No data de Boer et al. 

2000b 

Eels Netherlands <20–1,400 No data No data <50–1,700 No data de Boer 1990 

Osprey Sweden 1,800 140 200 2,140 No data Sellström et al. 

1993 

Pike Lake Bolmen, 

Sweden 

65 42 19 130 No data Kierkegaard et 

al. 1993 

Pike Dalslands canal, 

Sweden 

94–98 60–79 25–36 180–210 No data Sellström et al. 

1993 

Pike River Viskan, 

Sweden, 

upstream and 

downstream 

<46–2,000 <37–

1,600 

<14–

1,000 

<130–

4,600 

Trace Sellström et al. 

1998a 

Several fish 

species 

Germany No data No data No data 19–983* No data Krüger 1988 

Trout Dalslands canal, 

Sweden 

120–460 130–590 33–150 280–1,200 No data Sellström et al. 

1993 

Whitefish Lake 

Storvindeln, 

Sweden 

15 7.2 3.9 26 No data Sellström et al. 

1993 

Whitefish Lake Geneva, 

Switzerland 

26 13 2.5 44* No data Zennegg et al. 

2003 

Whitefish Lake Greifen, 

Switzerland 

96 52 9.1 165* No data Zennegg et al. 

2003 

Whitefish Lake Biel, 

Switzerland 

75.9 39 7.1 128* No data Zennegg et al. 

2003 

Whitefish Lake Lucerne, 

Switzerland 

56 46 10 121* No data Zennegg et al. 

2003 

Whitefish Lake Zürich, 

Switzerland 

56 25 4.5 89* No data Zennegg et al. 

2003 

Whitefish Lake Nauchatel 

Switzerland 

41 20 4.0 68* No data Zennegg et al. 

2003 

Whitefish Lake Constance, 

Switzerland 

32 15 2.9 52* No data Zennegg et al. 

2003 

Whitefish Lake Thun, 

Switzerland 

19 12 2.5 36* No data Zennegg et al. 

2003 

a

ΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an 

asterisk (*). 

Sources: de Wit 2002; Zennegg et al. 2003 

dw = dry weight; fw = fresh weight 
PBBs and PBDEs 432 

8. POTENTIAL FOR HUMAN EXPOSURE 

in pike and trout are of the same order of magnitude as in the Arctic char, indicating the spread of PBDEs 

from diffuse sources (de Wit 2002). In 1979 and 1980, high levels of tri- to hexaBDEs (range, 950–

27,000 ng/g lipid weight in muscle tissues) were measured in fish sampled along a river in Sweden 

(Viskan) where numerous textile industries are located (Andersson and Blomkvist 1981). These textile 

industries have used PBDEs in the production of textiles. BDE 47 was the predominant congener at 70–

80% of the total PBDEs. In 1977, the PBDEs were not detected in fish sampled at the same sites. The 

elevated levels of BDE 47, BDE 99, and BDE 100 were later confirmed in a follow-up study in which 

fish were caught from approximately the same locations (Sellström et al. 1993). In the current study, 

BDE 47 was the predominant congener at 65–96% of total PBDEs. Several fish species were sampled 

(pike, perch, bream, eel, tench, and sea trout) in these studies. In 1995, fresh samples of pike and 

sediments were collected at four of eight sites along River Viskan in order to search for point sources of 

contaminants. The combined concentrations of BDE 47, BDE 99, and BDE 100 ranged from not detected 

to 4,600 ng/g lipid weight, with BDE 47 being the predominant congener (50–90% of total). DecaBDE 

(BDE 209) was found in a few fish at trace amounts. The lowest levels of the PBDEs were found 

upstream of the industries. The concentrations of PBDEs increased further downstream as more 

industries were passed (Sellström et al. 1998a). Levels of BDE 47 ranged from <20 to 1,700 ng/g lipid in 

eels (Anguilla anguilla) from Dutch rivers and lakes (at 10 locations); BDE 47 comprised 70% of the total 

PBDEs (de Boer 1990). Bream (Abramais brama) sampled from several sites in the Netherlands had 

concentrations of BDE 47 ranging from 0.2 to 130 ng/g dry weight (de Boer et al. 2000). BDE 99 was 

below the detection limits. BDE 153 ranged from <0.04 to 4.1 ng/g dry weight. Allchin et al. (1999) 

conducted a study of PBDEs in plaice (Pleuronectes platessa), flounder (Platichys flesus), and dab 

(Limanada limanada) collected in the estuaries of rivers in the United Kingdom. Suspected sources of 

PBDEs in the estuaries include a manufacturer of pentaBDE and octaBDE, several industries using 

pentaBDE, and several landfills receiving wastes suspected to contain PBDEs. Levels of BDE 47, 

BDE 99, pentaBDE (as technical mixture DE-71), and octaBDE (as technical mixture DE-79) in fish 

ranged from not detected to 9,500 ng/g lipid weight, not detected to 370 ng/g lipid weight, 47 to 

1,200 ng/g lipid weight, and not detected to 1,200 ng/g lipid weight. The highest levels were at Tees Bay 

downstream from a manufacturing plant on the River Tees. These results are similar to the situation 

found in Sweden along the River Viskan (Andersson and Blomkvist 1981; Sellström et al. 1993). 

Freshwater mussels (Dreissena polymorpha) were collected at several locations in the Netherlands and 

analyzed for BDE 47, BDE 99, BDE 153, and BDE 209) (de Boer et al. 2000). Concentration ranges for 

the congeners were 0.7–17, 0.4–11, and <0.1–1.5 ng/g dry weight for BDE 47, BDE 99, and BDE 153, 

respectively; BDE 209 was below the detection limit. 
PBBs and PBDEs 433 

8. POTENTIAL FOR HUMAN EXPOSURE 

Saltwater Fish. No identifiable temporal trends were found for PDBE levels in marine aquatic species. 

Spatial trends show higher levels of lower brominated BDE congeners found near human populated areas. 

The congener profiles show the highest levels for 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47). The 

levels of several PBDEs in marine aquatic species are summarized in Table 8-9. In the year 2000, sole 

liver collected from five sites along the Canadian west coast (Crofton, Bamfield, Kitimat, Trincomali, and 

Vancouver) were analyzed for 14 BDE congeners (Ikonomou et al. 2002b); the total PBDE 

concentrations were 64–340 ng/g lipid while the three highest congener concentrations were 27–160 ng/g 

lipid (BDE 47), 8.5–54 ng/g lipid (2,2’,4,4’,6-pentabromodiphenyl ether [BDE 100]), and 9.5–46 ng/g 

lipid (2,2’,4,4’,5-pentabromodiphenyl ether [BDE 99]), respectively. The highest levels were found in 

sole samples collected near Vancouver, Canada. DecaBDE was not detected in these samples at the level 

of procedural blank. Farmed salmon collected at two locations in Canada were analyzed for PBDE 

congeners (Easton et al. 2002). Forty-one congeners were detected with BDE 47 at the highest level 

(690 and 2,600 ng/g wet weight) followed by BDE 99 and BDE 100; total BDE congener levels were 

1,188 and 4,147 ng/g wet weight for the two samples. Likewise, wild salmon from four locations in 

Canada were analyzed for BDE congeners. Levels were a factor of 10 lower for these samples compared 

to farmed salmon samples. The total PBDE concentration for the 41 detected congeners ranged from 

38.7 to 485.2 ng/g wet weight. The concentration of the highest congener, BDE 47, ranged from 29 to 

280 ng/g wet weight (Easton et al. 2002). PBDE concentrations in skipjack tuna from Asian offshore 

waters, off-Seychelles, off-Brazil, and open seas were determined for samples collected during 1996–

2001 (Ueno et al. 2003). The concentration of total BDEs in muscles tissues ranged from not detected 

(<0.05 ng/g lipid) to 53 ng/g lipid. The concentration of the highest congener in muscle tissues, BDE 47, 

ranged from <0.1 to 15 ng/g lipid. BDE 99, BDE 100, BDE 153, and BDE 154 also were detected; 

BDE 209 was below the detection limit (<5.0 ng/g lipid) for these samples. Samples collected off the 

coast of the Seychelles (relatively pristine area) did not have detectable levels of any PBDEs, while 

samples collected in industrial areas of southeast Asia had the highest. Fall-caught herring (Clupea 

harengus) muscle from five sites along the Swedish coast was analyzed for BDE 47, BDE 99, and 

BDE 100; the combined concentration of these three congeners ranged from 17 to 61 ng/g lipid, with 

BDE 47 being the dominate congener (Sellström et al. 1993). Likewise, the concentration of BDE 47 in 

Baltic herring ranged from 3.2 to 27 ng/g lipid in different age groups; the combined concentration of 

BDE 47, BDE 99, and BDE 100 ranged from 3.2 to 32 ng/g lipid (Haglund et al. 1997); 2-year-old 

herring had the lowest levels and 5-year-old herring had the highest levels. Similarly, Strandman et al. 

(1999) observed increasing concentrations with age of BDE 47, BDE 99, and BDE 153 in Baltic sprat 

(Sprattus sprattus, age 3–13 years). However, this trend was not evident for herring. BDE 47 was the 
PBBs and PBDEs 434 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-9. Concentrations (ng/g Lipid Weight) of Several PBDEs in Marine 

Aquatic Species 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa

 BDE 209 Reference 

Farmed 

Salmon 

Canada 690; 

2,600 

(ww) 

140; 390 

(ww) 

130; 470 

(ww) 

1,187; 

4,147 (ww)

No data Easton et al. 2002 

Salmon 

(wild) 

Canada 29–280 

(ww) 

ND–97 

(ww) 

4.2–43 

(ww) 

38.7–485.2 

(ww) 

No data Easton et al. 2002 

Sole liver West coast, 

Canada 

27–160 9.5–46 8.5–54 64–340* ND Ikonomou et al. 

2002 

Skipjack 

tuna 

Seychelles, 

Indian 

Ocean 

<0.1 <0.05 <0.05 ND <5.0 Ueno et al. 2003 

Skipjack 

tuna 

East China 

Sea 

9.0–15 2.4–4.7 3.4–4.4 23–34 <5.0 Ueno et al. 2003 

Skipjack 

tuna 

Pacific 

Ocean 

2.9–7.9 0.18–3.0 0.56–2.1 5.8–21 <5.0 Ueno et al. 2003 

Herring Baltic Sea 19–38 7.8–17 3.4–6 30–61 No data de Wit 2002 

Sellström et al. 

1993 

Herring Baltic Sea 3.2–27 ND–2.9 1.3–1.9 3.2–32 No data Haglund et al. 1997

Herring Baltic Sea 7.6–24 4.3–3.9 No data 12.9–28.3* No data Strandman et al. 

1999 

Herring Baltic Sea 6.3 0.6 0.8 12* No data Burreau et al. 1999

Herring Kattegatt, 

Sweden 

12 3.4 1.6 17 No data de Wit 2002; 

Sellström et al. 

1993 

Herring North Sea 8.4–100 No data No data No data No data de Boer 1990 

Sprat 

(different 

age groups) 

Baltic Sea 17.5–

140.8 

1.9–9.5 No data 21–149* No data Strandman et al. 

1999 

Sprat Baltic Sea 4.3 0.7 0.8 8.4* No data Burreau et al. 1999

Cod liver North Sea 170 No data No data 1.9–360 No data de Boer 1989 

Salmon Baltic Sea 167 52 44 220 No data Haglund et al. 1997

Salmon Baltic Sea 190 52 46 290 No data Asplund et al. 

1999b 

Salmon Baltic Sea 46 7.3 6.4 86* No data Burreau et al. 1999

Several fish 

species 

Japan No data No data No data 0.1–17* No data Watanabe et al. 

1987 

Yellowfin 

tuna 

Japan 0.5 0.4 0.25 1.9* No data Ohta et al. 2000 

Yellowtail Japan 17 4.5 4.0 30.5* No data Ohta et al. .2000 

Yellowtail 

(cultured) 

Japan 29 3.3 5.3 44* No data Ohta et al. 2000 

Salmon Japan 22 8.1 5.3 46* No data Ohta et al. 2000 
PBBs and PBDEs 435 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-9. Concentrations (ng/g Lipid Weight) of Several PBDEs in Marine 

Aquatic Species 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa

 BDE 209 Reference 

Several 

flatfish 

Seven river 

estuaries, 

Great Britain 

73–9,500 16–790 No data No data ND Allchin et al. 1999 

Flounder Netherlands, 

several sites 

0.6–20 

(dw) 

<0.01–4.6 No data No data No data de Boer et al. 

2000b 

a

ΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an 

asterisk (*). 

Source: Easton et al. 2002; de Wit 2002; Ikonomou et al. 2002; Ueno et al. 2003 
PBBs and PBDEs 436 

8. POTENTIAL FOR HUMAN EXPOSURE 

primary congener with levels ranging from 7.6 to 24 ng/g lipid weight for 1–3-year-old sprat, 17–

140 ng/g lipid weight for 3- to 13-year-old sprat, and 7.6–24 ng/g lipid weight in herring. The 

concentrations of BDE 47, BDE 99, and BDE 100 in whole-body composites of herring were 6.21, 0.62, 

and 0.81 ng/g lipid, respectively; in sprat, the concentrations were 4.32, 0.71, and 0.80 ng/g lipid, 

respectively (Burreau et al. 1999). Baltic sea herring had similar levels of BDE 47 (46.3 ng/g lipid) 

compared to 8.4–100 ng/g lipid of BDE 47 found by de Boer (1990) for herring collected from three 

regions of the North Sea. BDE 47, BDE 99, and BDE 153 concentrations in Baltic salmon (Salmo salar) 

muscle were 167, 52, and 4.2 ng/g lipid, respectively (Haglund et al. 1997). BDE 47, BDE 99, and 

BDE 100 levels were 47, 7.2, and 6.3 ng/g lipid, respectively, in whole-body composites (Burreau et al. 

1999). In another study, the levels of BDE 47, BDE 99, and BDE 100 were determined in muscle, ripe 

eggs, and blood plasma from Baltic salmon (Asplund et al. 1999a). The mean concentrations of PBDEs 

in tissues from Baltic salmon (ng/g lipid weight) were as follows: BDE 47 (muscle, 190; ripe eggs, 64; 

blood, 190), BDE 99 (muscle, 52; ripe eggs, 16; blood, 55), and BDE 100 (muscle, 46; ripe eggs, 18; 

blood, 59). Cod (Gadus morhua) liver samples at three locations of the North Sea had combined levels of 

BDE 47 and BDE 99 of 1.9–360 ng/g lipid (de Boer 1989). BDE concentrations in flounder were 0.6–

20 ng/g dry weight for BDE 47 and <0.01–4.6 ng/g dry weight for BDE 99 from several sites in the 

Netherlands (de Boer et al. 2000). Concentrations of BDE 153 and BDE 209 were below the detection 

limit. In 1996, de Boer et al. (2001) measured the levels of two BDE congeners in flounder liver samples 

from the Amsterdam and Rotterdam harbors, and off the Dutch coast; BDE 47 and BDE 99 ranged from 

15 to 280 and from <2 to 24 ng/g lipid weight, respectively. Olsson et al. (1999) detected BDE 47 in 

perch (Perca fluviatilis) from Latvia in a study examining environmental contamination in coastal areas 

of the former Soviet Union; the concentration of BDE 47 ranged from 6.4 to 10 ng/g lipid weight in the 

perch. 

Watanabe et al. (1987) detected PBDEs in numerous marine fish and shell fish in Japan. TetraBDE and 

pentaBDE levels ranged from 0.1 and 17 ng/g fresh weight, with tetraBDE being the major congener. 

decaBDE was also detected in a mussel sample from Osaka Bay (at 1.4 μg/kg wet weight). Recently, 

Japanese market fish were analyzed for PBDEs. The highest combined PBDE levels (2,4,4’-triBDE 

[BDE 28], 2,2’,4,4’-tetraBDE [BDE 47], 2,3’,4,4’-tetraBDE [BDE 66], 2,2’,4,4’,5-pentaBDE [BDE 99], 

2,2’,4,4’,6-pentaBDE [BDE 100], 2,2’,4,4’,5,5’-hexaBDE [BDE 153], and 2,2’,4,4’,5,6’-hexaBDE 

[BDE 154]) were in salmon, cultured yellowtail, and wild yellowtail muscle (46, 44, and 30.5 ng/g lipid 

weight, respectively) and lowest levels in yellowfin tuna (1.9 ng/g lipid weight) (Ohta et al. 2000). 

BDE 47 was major congener in all samples. In another study, several fish species from Japan were 

analyzed for 15 BDE congeners (Hori et al. 2000). The PBDE levels ranged from 0.00136 to 2.1 ng/g 
PBBs and PBDEs 437 

8. POTENTIAL FOR HUMAN EXPOSURE 

fresh weight, with BDE 47 as the predominant congener. Seven species of marine fish (conger eel, 

flounder, gray mullet, horse mackerel, red sea bream, sea bass, and yellowtail) were collected from the 

Inland Seas near Seto, Japan (Akutsu et al. 2001). Seven PBDEs (BDE 28, BDE 47, 2,3’,4,4’-tetraBDE 

[BDE 66], BDE 99, BDE 100, BDE 153, and BDE 154) were detected in all samples with BDE 47 being 

the most abundant congener. Levels of total PBDEs in gray mullets and yellowtails were 63 and 15 ng/g 

lipid weight, respectively. 

Marine Aquatic Organisms. Marine mussels (Mytilus edulis) collected at several locations in the 

Netherlands and analyzed for 2,2’,4,4’-tetraBDE (BDE 47), 2,2’,4,4’,5-pentaBDE (BDE 99), 

2,2’,4,4’,5,5’-hexaBDE (BDE 153), and decaBDE (BDE 209) (de Boer et al. 2000). Concentrations of 

BDE 47 and BDE 99 ranged from 0.9 to 4.3 ng/g dry weight and from 0.3 to 1.6 ng/g dry weight, 

respectively. BDE 153 and BDE 209 were not detected. Di- to heptaBDE were analyzed for in 

hepatopancreas samples from Dungenes crab from several sites on the Strait of Georgia, British 

Columbia, Canada (Ikonomou et al. 1999). The primary congener detected was BDE 47. The combined 

concentration of BDE 47 and BDE 99 was approximately 100–350 ng/g lipid weight. 

Marine Animals. In marine animals, temporal trends show increasing levels of lower brominated BDE 

congeners with higher levels found near human-populated areas. In all marine animal studies, the 

congeners profile show the highest levels for BDE 47. The concentrations of several PBDEs in marine 

animals are summarized in Table 8-10. PBDEs have been detected in several species of seal from several 

different sites. In San Francisco Bay, California, the concentrations of PBDEs in harbor seals have 

increased dramatically over the past decade, with current levels among the highest reported for this 

species (She et al. 2002). The concentration of total PBDEs (sum of BDE 47, 2,2’,4,4’,5-pentaBDE 

[BDE 99], 2,2’,4,4’,6-pentaBDE [BDE 100], 2,2’,4,4’,5,5’-hexaBDE [BDE 153], and 

2,2’,4,4’,5,6’-hexaBDE [BDE 154]) in harbor seal blubber increased by over a factor of 50 from a 

concentration of 88 ng/g lipid for species samples collected in 1988 to a concentration of 2,985–

8,325 ng/g lipid for species samples collected in 1998. The highest concentrations reported were for 

BDE 47, which increased from 45.6 ng/g lipid for blubber samples in 1989 to 2,343–6,682 ng/g lipid for 

blubber samples collected in 1998. The dominance of the tetraBDE congeners over other congeners may 

indicate that tetraBDEs bioaccumulate more than the higher brominated congners (She et al. 2002). In the 

Baltic Sea, female grey seals (Halichoerus grypus) sampled in 1979–1985 contained 730 ng PBDE/g lipid 

in their blubber (sum of BDE 47, BDE 99, and BDE 100) (Jansson et al. 1993); male grey seals had 

280 ng PBDE/g lipid weight (Andersson and Wartanian 1992). Male ringed seals (Pusa hispida) from the 

Baltic Sea had 320 ng PBDE/g lipid weight (Andersson and Wartanian 1992). Baltic gray and ringed seal 
PBBs and PBDEs 438 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-10. Concentrations (ng/g Lipid Weight) of Several PBDEs in 

Marine Animals 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa

 BDE 209 Reference 

Bottlenose 

dolphin 

Gulf of Mexico No data No data No data 8,000 No data Kuehl and Haebler 

1995 

Harbor seal San Francisco 

Bay, California 

46–

6,682 

17–303 No data No data No data She et al. 2000 

Harbor seal 

(blubber) 

San Francisco 

Bay, California 

1,304 112 87.1 1,730* No data She et al. 2002 

Herring Gull 

Eggs 

Lake Superior, 

United States 

253–323 

(fw) 

202–284 

(fw) 

83.6–113 

(fw) 

664–887 

(fw)* 

No data Norstrom et al. 2002

Herring Gull 

Eggs 

Lake Michigan, 

United States 

522–602 

(fw) 

323–459 

(fw) 

167–203 

(fw) 

1,366–

1,400 (fw)*

No data Norstrom et al. 2002

Herring Gull 

Eggs 

Lake Huron, 

United States 

and Canada 

146–291 

(fw) 

74.6–161 

(fw) 

37.3–89.5 

(fw) 

308–652 

(fw)* 

No data Norstrom et al. 2002

Herring Gull 

Eggs 

Detroit River, 

United States 

322 (fw) 130 (fw) 92.6 (fw) 639 (fw)* No data Norstrom et al. 2002

Herring Gull 

Eggs 

Lake Erie, 

United States 

70–163 

(fw) 

52–55.9 

(fw) 

24.6–51.8 

(fw) 

192–340 

(fw)* 

No data Norstrom et al. 2002

Herring Gull 

Eggs 

Niagara River, 

United States 

168 (fw) 111 (fw) 53 (fw) 432 (fw)* No data Norstrom et al. 2002

Herring Gull 

Eggs 

Lake Ontario, 

Canada 

220–401 

(fw) 

113–322 

(fw) 

66.5–102 

(fw) 

530–1,003 

(fw)* 

No data Norstrom et al. 2002

Herring Gull 

Eggs 

St Lawrence 

River, United 

States 

220 (fw) 89.8 (fw) 56.6 (fw) 453 (fw)* No data Norstrom et al. 2002

Beluga 

whale 

Canadian Arctic No data No data No data 81–160* No data Alaee et al. 1999 

Beluga 

whale 

Southeast 

Baffin, Canada 

10 0.9 1.6 15* No data Stern and Ikonomou 

2000 

Bottlenose 

dolphin 

South Atlantic 

Ocean 

No data No data No data 180–220 No data Kuehl et al. 1991 

Brunnich’s 

guillemot 

Svalbard, 

Sweden 

No data No data No data 130 No data Jansson et al. 1987 

Cormorant England, 

United 

Kingdom 

170–

3,500 

50–250 50–1,500 300–6,400* No data Allchin et al. 2000 

Cormorant 

liver 

Rhine delta, 

Germany 

No data No data No data 28,000 (fw) No data de Boer 1990 

Galaucous 

gull 

Bear Island, 

Norway (Arctic) 

290–634 160 No data No data No data de Wit 2002 

Grey seal Baltic Sea 650 40 38 730 No data de Wit 2002; 

Sellström et al. 1993

Grey seal Baltic Sea 308 54 57 419 No data Haglund et al. 1997 

Grey seal Baltic Sea No data No data No data 208 No data Andersson and 

Wartanian 1992 
PBBs and PBDEs 439 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-10. Concentrations (ng/g Lipid Weight) of Several PBDEs in 

Marine Animals 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa

 BDE 209 Reference 

Harbor 

porpoise 

Britsh 

Columbia, 

Canada 

50–

1,200 

No data No data 350–2,300* No data Ikononmou et al. 

2000b 

Harbor 

porpoise 

England and 

Wales, United 

Kingdom 

227–

6,790 

No data No data 440–7,670 No data Law et al. 2000 

Harbor seal Baltic Sea No data No data No data 90 No data Jansson et al. 1987 

Harbor seal Skagerrak, 

Norway and 

Sweden 

No data No data No data 230 No data Andersson and 

Wartanian 1992 

Harbor seal North Sea 390–

4,900 

42–660 25–450 600–6,000 No data de Boer et al. 1998 

Long-finned 

pilot whale 

Faeroe Islands 410–

1,780 

160–600 87–280 843–3,160* No data Lindström et al. 1999

Long-finned 

pilot whale 

Faeroe Islands 66–860 24–170 12–98 126–1,250* No data van Bavel et al. 1999

Minke whale Netherlands 630 160 79 870 No data de Boer et al. 1998 

Ringed seal Baltic sea 256 33 61 350 No data Haglund et al. 1997 

Ringed seal Baltic sea No data No data No data 320 No data Andersson and 

Wartanian 1992 

Ringed seal Svalbard, 

Sweden 

47 1.7 2.3 51 No data de Wit 2002; 

Sellström et al. 1993

Ringed seal Canadian Arctic No data No data No data 25.8–50* No data Alaee et al. 1999 

Ringed seal Holman Island, 

Northwest Territories, Canada 

2.8 No data No data 2.4–4.9* No data Ikonomou et al. 2000

Sperm whale Netherlands 130–250 32–64 21–35 187–349 No data de Boer et al. 1998 

Whitebeaked 

dolphin 

Netherlands 5,500 1,000 1,200 7,700 No data de Boer et al. 1998 

a

ΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an 

asterisk (*). 

Sources: de Wit 2002; Norstrom et al. 2002; She et al. 2002 

fw = fresh weight 
PBBs and PBDEs 440 

8. POTENTIAL FOR HUMAN EXPOSURE 

blubber sampled between 1981 and 1988 contained 419 and 350 ng PBDEs/g lipid (total of BDE 47, 

BDE 99, and BDE 100), respectively (Haglund et al. 1997). In 1981, female ringed seals from Svalbard 

in the Swedish Arctic contained 40–51 ng PBDEs/g lipid in blubber (Jansson et al. 1987, 1993; Sellström 

et al. 1993). Higher levels of PBDEs are generally evident in Baltic Sea ringed seals (320–350 ng/g lipid) 

(Andersson and Wartanian 1992; Haglund et al. 1997) compared to Arctic ringed seals (26–51 ng/g lipid) 

(Alaee et al. 1999; Jansson et al. 1987). The level of PBDEs in harbor seals from Skagerrak on the 

Swedish west coast was 230 ng PBDE/g lipid (Andersson and Wartanian 1992). She et al. (2000) 

analyzed the concentration of BDE 47, BDE 99, and BDE 153 in harbor seal from the San Francisco Bay 

area (She et al. 2000). Mean concentrations for BDE 47, BDE 99, and BDE 153 were 1,124, 107, and 

50 ng/g lipid weight, respectively. Alaee et al. (1999) found that ringed seal from the Canadian Arctic 

had mean PBDE concentrations (sum of di- to hexaBDEs) of 25.8 ng/g lipid weight (females) and 

50.0 ng/g lipid weight (males). The lower levels in female seals suggest that PBDEs are transferred to 

young through breast milk. On Holman Island, Northwest Territory, Canada (Arctic) in 1996, ringed seal 

had total PBDE concentrations of 2.4–4.9 ng/g lipid for males. The levels of PBDEs were found to 

increase with age (Ikonomou et al. 2000). In a temporal trend study, archived samples of blubber from 

ringed seals from Holman Island, Northwest Territory, Canada were analyzed for PBDE levels. The 

concentration of PBDE in samples collected between 1981 and 1996 increased from approximately 

0.3 ng/g lipid weight in 1981 to 3.6 ng/g lipid weight in 1996 (Ikonomou et al. 2000). 

The levels of PBDEs have recently been determined in harbor porpoises (Phocaena phocaena) from 

British Columbia, Canada (Ikonomou et al. 2000) and from the coasts of England and Wales (Law et al. 

2000). In British Colombia (Canada) samples, the total PBDE levels (sum of tri- to hepta-congeners) 

were 350–2,300 ng/g lipid weight; 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47) was found at the highest 

levels in these samples (range, 50–1,200 ng/g lipid weight) (Ikononmou et al. 2000). Concentrations of 

total PBDEs (sum of 13 congeners) along the coast of England and Wales, ranged from 450 to 7,670 ng/g 

lipid weight, with BDE 47 levels ranging from 227 to 6,790 ng/g lipid weight (Law et al. 2000). 

During a mass mortality event on the south Atlantic coast in 1987–1988, blubber samples were collected 

from three bottlenose dolphins (Tursiops truncatus); these samples contained 180–220 ng PBDEs/g lipid 

(Kuehl et al. 1991). Blubber samples, taken from stranded bottlenose dolphins from several locations 

around the Gulf of Mexico in 1990, contained 3,110 ng PBDEs/g lipid (Kuehl and Haebler 1995). On the 

Dutch coast in early 1998, de Boer et al. (1998) found PBDEs in blubber of one whitebeaked dolphin 

(Lagenorhynchus albirostris); the levels of 2,2’4,4’-tetraBDE (BDE 47), 2,2’,4,4’,5-pentaBDE (BDE 99), 

and 2,2’,4,4’,6-pentaBDE (BDE 100) were 5,500, 1,000, and 1,200 ng/g lipid weight, respectively. 
PBBs and PBDEs 441 

8. POTENTIAL FOR HUMAN EXPOSURE 

The concentration of 19 PBDEs was determined in long-finned pilot whale (Globicephala melas) from 

the Faeroe Islands in the north Atlantic (Lindström et al. 1999). Young males and females had the highest 

levels, ranging from 3,000 to 3,160 ng/g lipid; lower levels were observed for both adult females (840–

1,050 ng/g lipid) and males (1,610 ng/g lipid). The predominant isomers in all samples were BDE 47 and 

2,2’,4,4’,5-pentaBDE (BDE 99), accounting for 70% of the sum of 19 congeners. van Bavel et al. (1999) 

also studied the levels of PBDEs in long-finned pilot whales. They observed a similar trend with young 

animals having higher PBDE concentrations (740 ng/g lipid weight) and adult animals having lower 

levels (females, 230 ng/g lipid; males, 540 ng/g lipid). In Beluga whales sampled in 1997 from southeast 

Baffin (Cumberland Sound), the levels of total PBDEs and BDE 47 were 15 and 10 ng/g lipid weight, 

respectively (Stern and Ikonomou 2000). Between 1982 and 1997, total PBDE concentrations in archived 

blubber samples of beluga whales from southeast Baffin Canada increased significantly. For this time 

period, BDE 47, BDE 99, 2,2’,4,4’,6-pentaBDE (BDE 100), and 2,2’,4,4’,5,6’-hexaBDE (BDE 154), and 

total PBDEs increased by factors of 6.5, 10.3, 7.9, 30.6, and 6.8, respectively (Stern and Ikonomou 2000). 

Three sperm whales (Pyseter macrocephalus) and one minke whale (Balaenaoptera acutorostrata) found 

stranded on the Dutch coast in early 1998 were analyzed for PBDEs (de Boer et al. 1998a, 1998b). 

Exposure to PBDEs for these animals occurred in the deep Atlantic through the food web. The 

concentrations of PPBEs in these marine mammals were as follows: sperm whale (BDE 47, 130–250 ng/g 

lipid weight; BDE 99, 32–64 ng/g lipid weight; and BDE 100, 21–35 ng/g lipid weight) and minke whale 

(BDE 47, 630 ng/g lipid weight; BDE 99, 160 ng/g lipid weight; BDE 100, 79 ng/g lipid weight); 

BDE 209 (decaBDE) was below detection limits in all samples. 

Marine Birds. Increasing levels of PBDEs have been found in marine birds and eggs, with 2,2’4,4’-tetraBDE (BDE 47) found at the highest levels. Di- and triBDE have been detected, but not quantified, in 

black skimmer (Rynchops nigra) tissues and eggs in the United States (Stafford 1983). In 2000, herring 

gull eggs collected from 15 locations around the Great Lakes (United States and Canada) were pooled and 

analyzed for PBDEs (Norstrom et al. 2002). A total of 25 di- to hepta-BDE congeners were identified in 

herring gull throught the Great Lakes system. No mono-, octa-, nona-, or decaBDEs were found at the 

detection limit of the analysis (0.01–0.05 ng/g wet weight). Seven congeners, 2,4,4’-triBDE (BDE 28), 

BDE 47, 2,2’,4,4’,5-pentaBDE (BDE 99), 2,2’,4,4’,6-pentaBDE (BDE 100), 2,2’,4,4’,5,5’-hexaBDE 

(BDE 153), 2,2’,4,4’,5,6’-hexaBDE (BDE 154), and 2,2’,3,4,4’,6,6’-heptaBDE (BDE 184) constituted 

97.5% of total PBDEs (192–1,400 ng/g wet weight). BDE 47 was the dominant congener (70–602 ng/g 

wet weight) followed by BDE 99 (52–459 ng/g wet weight). The highest concentrations (1,003–

1,400 ng/g wet weight) were found in two Lake Michigan colonies and in Toronto Harbor, Lake Ontario 
PBBs and PBDEs 442 

8. POTENTIAL FOR HUMAN EXPOSURE 

(Norstrom et al. 2002). Muscle tissues from ospreys (Pandion haliaetus), found dead at various locations 

around Sweden, were pooled and analyzed for PBDEs (Jansson et al. 1993; Sellström et al. 1993). The 

ospreys’ diet was freshwater fish. The combined concentration of BDE 47, BDE 99, and BDE 100 was 

2,100 ng/g lipid in samples collected between 1982 and 1986; BDE 47 was the primary congener (86%) 

in these samples (n=35). High concentrations of PBDEs may reflect biomagnification and/or fish 

consumption along their migratory routes. The concentrations of PBDEs in common guillemots (Uria 

aalge) collected in 1979–1981 from the Baltic and North Seas were 370 and 80 ng/g lipid, respectively 

(Jansson et al. 1987). As part of the Swedish National Environmental Monitoring Program, guillemot 

eggs (St. Karlsö, Baltic Sea) are collected yearly and placed in the Swedish Natural History Museum’s 

Environmental Specimens Bank. The concentrations of BDE 47, BDE 99, and BDE 100 in pooled egg 

samples from the specimen bank showed a significant increase from 1969 to the beginning of the 1990s, 

with highs of 1,100 ng/g for BDE 47 in 1984 and 190 ng/g for BDE 99 in 1990 (Sellström 1999; 

Sellström et al. 1993). Between 1992 and 1997, PBDE levels started to decrease statistically. In 1997, 

the PBDE level (sum of BDE 47, BDE 99, and BDE 100) was 190 ng/g lipid, with BDE 47 as the 

predominant congener. 

Human Body Tissues and Fluids. The quantitative determination of the concentrations of PBDEs in 

body tissues and fluids is important in determining the human body burden of these chemicals. 

Increasing levels of lower brominated PBDEs have been measured in blood and breast milk in temporal 

trend studies. Individuals who consumed fish had a somewhat higher concentration of total PBDEs in 

body fluids compared to individuals who ate less fish. 

Tables 8-11, 8-12, and 8-13 summarize the concentrations of PBDEs found in blood (serum), adipose 

tissue, breast milk, and other body tissues or fluids, respectively. Levels of PBDEs in body tissues and 

fluids from individuals living in the United States have recently been determined. These studies indicate 

that levels of lower brominated BDEs in body fluids are a factor of 10–100-fold higher for individuals 

living in the United States compared to individuals living in other regions of the world (e.g., Europe). 

Serum samples collected from 12 U.S. blood donors in 1988 were analyzed for PBDEs, and BDE 47, 

BDE 153, BDE 183, and BDE 209 were detected (Patterson et al. 2000; Sjödin et al. 2001). 

Concentrations of these congeners were similar to those found in the Sjödin et al. (1999b) study for the 

control group. The median concentrations and ranges of BDE 47, BDE 153, BDE 183, BDE 209, and 

total PBDEs (sum of four congeners) were 0.63 (<0.4–24); 0.35 (0.08–2.0); 0.17 (0.09–1.3); <1 (<1–34); 

and 2.2 ng/g lipid weight, respectively (Sjödin et al. 2001). DecaBDE was found at levels above the limit 

of quantification (1 pmol/g lipid) in 5 of 12 serum samples (Patterson et al. 2000). Serum samples were 
PBBs and PBDEs 443 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-11. Concentrations (ng/g Lipid Weight) of Several PBDEs in Human 

Blood Samples 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa

 BDE 209 Reference 

Human 

blood 

San Francisco, 

California 

ND No data No data No data No data Petreas et al. 2002

Human 

blood 

San Francisco, 

California 

95 No data No data No data No data Petreas et al. 2002

Human 

blood 

San Francisco, 

California 

16.5 No data No data No data No data Petreas et al. 2003

Human 

blood 

United States (in 

1988) 

0.63 

(median) 

0.32 

(median) 

0.17 

(median) 

No data <0.1 Sjödin et al. 2001 

Maternal 

serum 

Indiana 28 (9.2–

310) 

5.7 (2.4–

68) 

4.2 (1.9–

110) 

37 (15–

580) 

No data Mazdai et al. 2003 

Fetal 

serum 

Indiana 25 (8.4–

210) 

7.1 (2.2–

54) 

4.1 (1.8–

91) 

39 (14–

460) 

No data Mazdai et al. 2003 

Human 

blood 

Sweden No data No data No data 2.1* No data Klasson Wehler et 

al. 1997 

Human 

blood 

Sweden, 

computer disassembly 

workers 

2.9 

(median) 

No data No data 26* 4.8 Sjödin et al. 1999a 

Human 

blood 

Sweden, 

cleaning 

personnel/office 

workers 

1.5–1.6 

(median) 

No data No data 3.3–4.1* <0.7 

(median) 

Sjödin et al. 1999a 

Human 

blood 

Sweden, high 

fish intake 

2.1 No data No data No data No data Bergman et al. 

1999; Sjödin et al. 

2000 

Human 

blood 

Sweden, no fish 

intake 

0.40 No data No data No data No data Bergman et al. 

1999; Sjödin et al. 

2000 

Maternal 

blood 

Sweden 0.83 (0.3–

5.1) 

0.19 

(<0.01–

1.43) 

0.17 

(<0.01–

0.52) 

2.07 (0.71–

8.39) 

No data Meironyte 

Guvenius et al. 

2003 

Cord 

blood 

Sweden 0.98 

(0.33–

3.28) 

0.07 0.07 0.46–4.28 No data Meironyte 

Guvenius et al. 

2003 

Human 

blood 

Germany 3.9 No data No data 5.6* No data Schröter-Kermani 

et al. 2000 

a

ΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an 

asterisk (*). 

Source: de Wit 2002; Sjödin et al. 2001 
PBBs and PBDEs 444 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-12. Concentrations (ng/g Lipid Weight) of Several PBDEs in Human 

Adipose Tissue Samples 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa BDE 209 Reference 

Human 

adipose 

tissue 

Northern 

California 

7.0–28 3.1–7.3 No data No data No data She et al. 2000 

Human 

adipose 

tissue 

San 

Francisco, 

California 

16.5 

(5.2–

196) 

No data No data No data No data Petreas et al. 2003 

Human 

adipose 

tissue 

United States No data No data No data No data ND–0.7 Cramer et al. 1990; 

Stanley et al. 1991 

Human 

adipose 

tissue 

Sweden 8.8 1.1 1.8 11.7 No data Haglung et al. 1997 

Human 

adipose 

tissue 

Sweden 3.8–16 No data No data No data No data Lindström et al. 1998 

Human 

adipose 

tissue 

Sweden 2.2 1.6 0.1 5* No data Meironyté Guvenius 

and Norén 1999 

Human 

adipose 

tissue 

Finland 7.3 2.3 No data 6.2–22* No data Strandman et al. 1999

Human 

adipose 

tissue 

Finland 1.20 0.26 0.09 No data No data Smeds and Saukko 

2003 

Human 

adipose 

tissue 

Spain 1.36 0.42 No data No data No data Meneses et al. 1999 

Human 

adipose 

tissue 

Japan 459 118 250 1,288 No data Choi et al. 2003 

a

ΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an 

asterisk (*). 

Source: de Wit 2002; Petreas et al. 2003 

ND = not detected 
PBBs and PBDEs 445 

8. POTENTIAL FOR HUMAN EXPOSURE 

Table 8-13. Concentrations (ng/g Lipid Weight) of Several PBDEs in Human 

Breast Milk Samples 

Sample 

type Location BDE 47 BDE 99 BDE 100 ΣPBDEsa

 BDE 209 Reference 

Human 

breast milk 

Texas 18.4 5.7 2.9 34.0 8.24 (max) Schecter et al. 

2003b 

Human 

breast milk 

Germany No data No data No data 0.6–11* No data Krüger 1988 

Human 

breast milk 

Uppsala 

County, 

Sweden 

2.35 0.62 0.38 4.01 No data 

Human 

breast milk 

Sweden 2.3 0.5 0.4 4* No data Norén and 

Meironyté 1998, 

2000 

Human 

breast milk 

Sweden 2.5 0.7 0.5 4.4* No data Darnerud et al. 

1998 

Human 

breast milk 

Finland 1.31 0.39 No data No data No data Strandman et al. 

2000 

Human 

breast milk 

Quebec and 

Ontario, 

Canada 

3.4 1.2 0.44 5.8* No data Ryan and Patry 

2000 

Human 

breast milk 

Maritimes, 

Canada 

No data No data No data 19* No data Ryan and Patry 

2000 

Human 

breast milk 

Quebec, 

Canada 

No data No data No data 18.8* No data Ryan and Patry 

2000 

Human 

breast milk 

Ontario, 

Canada 

No data No data No data 2.8* No data Ryan and Patry 

2000 

Human 

breast milk 

Prairies, 

Canada 

No data No data No data 5.7* No data Ryan and Patry 

2000 

Human 

breast milk 

Canada 

(wide area) 

No data No data No data 16.2 No data Ryan and Patry 

2000 

Human 

breast milk 

Japan 0.18–0.57 0.09–0.13 0.07–0.18 0.65–1.48* No data Ohta et al. 2000 

a

ΣPBDEs is the sum of BDE 47, BDE 99, and BDE 100, but if more congeners are included, this is marked with an 

asterisk (*). 
PBBs and PBDEs 446 

8. POTENTIAL FOR HUMAN EXPOSURE 

collected from a group of 50 Laotian immigrants (aged 19–40) participating in a reproductive outcome 

study in the San Francisco Bay area (Petreas et al. 2002). Participants were recruited and sampled in the 

late 1990s. The mean level of BDE 47 in serum was approximately 95 ng/g lipid. The contemporary 

samples were compared to serum samples taken from a group of over 400 women from the San Francisco 

Bay in the 1960s. Levels of BDE 47 in all archived samples were below the limit of detection. Recently, 

Petreas et al. (2003) expanded their investigation to include a diverse group of local women from the San 

Francisco Bay area sampled in the late 1990s. Their results confirmed earlier findings reported in Petreas 

et al. (2002). Mean concentrations of BDE 47 in serum samples taken from California women ranged 

from 5 to 510 ng/g lipid, with a median (16.5 ng/g lipid) 3–10 times higher than those reported from 

Europe (Petreas et al. 2003). In 2001, Mazdai et al. (2003) determined the concentration of six PBDE 

congeners (BDE 47, 2,2’,4,4’,5-pentaBDE [BDE 99], 2,2’,4,4’,6-pentaBDE [BDE 100], BDE 153, 

2,2’,4,4’,5,6’-hexaBDE [BDE 154], and BDE 183) and total PBDEs in maternal and fetal blood samples 

taken from subjects in Indianapolis, Indiana. Median levels of total PBDE (sum of six congeners) were 

39 and 37 ng/g lipid for fetal and maternal serum, respectively. BDE 47 was the predominant congener 

reported at median concentrations of 25 and 28 ng/g lipid for fetal and maternal serum samples, 

respectively. When compared with serum PBDE levels for a similar population of Swedish mothers and 

newborns, the levels for the Indiana population were 20- to 69-fold higher for maternal blood and 30- to 

106-fold higher for fetal blood. In fact, the median blood levels for this study were comparable to 

Swedish workers considered to have direct work-related exposures. These observations indicated that 

women in some areas of North America are exposed to much higher levels of lower brominated BDEs 

(i.e., BDE 47) than are European women. In general, the PBDE congener profile found in human serum 

was similar to that detected in environmental samples, except that there was an apparent decrease in the 

proportion of BDE 99. BDE 183 was detected in <17% of the samples even though it is the primary 

congener in octaBDE commercial mixtures (Mazdai et al. 2003). 

Six PBDE congeners (2,4,4’-triBDE [BDE 28], 2,2’,4,4’-tetraBDE [BDE 47], 2,3’,4,4’-tetraBDE 

[BDE 66], 2,2’,4,4’,5-pentaBDE [BDE 99], 2,2’,4,4’,6-pentaBDE [BDE 100], and 2,2’,4,4’,5,5’-hexaBDE [BDE 153]) were quantified in 40 human blood-plasma samples from Sweden. The highest 

concentrations in plasma were for BDE 47 and BDE 99; these congeners made up 70% of the total PBDE 

concentration. The mean concentration of total PBDEs were 2.1±1.4 ng/g lipid weight (Klasson Wehler 

et al. 1997). Whole-blood samples from a German environmental specimen bank, collected in 1985, 

1990, 1995, and 1999, contained measurable quantities of BDE 28, BDE 47, BDE 66, 2,2’,3,4,4’-hexaBDE (BDE 85), BDE 99, BDE 100, BDE 153, and 2,2’,4,4’,5,6’-hexaBDE (BDE 154). An increasing 

temporal trend was also observed; the mean total PBDE concentration (sum of eight congeners) increased 
PBBs and PBDEs 447 

8. POTENTIAL FOR HUMAN EXPOSURE 

from 3.9 ng/g lipid weight in 1985 to 5.6 ng/g lipid weight in 1999. For the 1999 sample, BDE 47 was 

the major congener found, with a mean concentration of 3.9 ng/g lipid weight. The total PBDE 

concentrations were significantly lower in female blood samples (Schröter-Kermani et al. 2000). In a 

study of the influence of diet on concentrations of PBDEs, BDE 47 was measured in blood serum from 

persons with high fish intake and no fish intake (Bergman et al. 1999; Sjödin et al. 2000). High fish 

intake groups of Swedish and Latvian men had median BDE 47 concentrations of 2.2 and 2.4 ng/g lipid 

weight, respectively, whereas the no fish-intake groups had median concentrations of 0.4 and 0.26 ng/g 

lipid weight, respectively (Sjödin et al. 2000). 

2,2’,4,4’-TetraBDE (BDE 47), 2,2’,4,4’,5,5’-hexaBDE (BDE 153), 2,2’,4,4’,5,6’-hexaBDE (BDE 154), 

2,2’,3,4,4’,5’,6-heptaBDE (BDE 183), and decaBDE (BDE 209) were measurable in blood plasma from 

three groups of workers (i.e., workers at a computer-disassembly plant, workers in a computerized office, 

and a control group) (Sjödin et al. 1999b). The median concentrations (sum of five congeners) were 

highest for the computer-disassembly plant workers (26 ng/g lipid weight); the office workers had a 

median concentration of 4.1 ng/g lipid weight and the control group had a median concentration of 

3.3 ng/g lipid weight. The congener patterns for the control group and office workers were similar, with 

BDE 47 having the highest levels. For the computer disassembly plant workers, the median 

concentrations of BDE 183, BDE 153, BDE 154, BDE 47, and BDE 209 were 7.8, 4.5, 1.2, 2.9, and 

4.8 ng/g lipid weight, respectively. Blood serum samples from 19 full-time computer technicians were 

analyzed (Hagmar et al. 2000). The serum concentrations of BDE 153, BDE 183, and BDE 209 in these 

samples were found to be approximately 5 times higher than the control and office workers in the Sjödin 

et al. (1999b) study. The median concentration for total PBDEs (for the sum of five congeners) was 

10.6 pmol/g (7.0 ng/g) lipid weight. The highest concentrations were of BDE 153. Two octaBDE 

congeners and one nonaBDE congener were also detected. Connections were observed between fish 

consumption and serum concentrations for congeners BDE 47, BDE 153, and BDE 183, and between 

worktime at the computer and congeners BDE 153 and BDE 183. 

DecaBDE, as well as hexa- through nonaBDE, has been found in composite samples from the 1987 

National Human Adipose Tissue Survey repository (Cramer at al 1990; Stanley et al. 1991). The 

concentrations ranged from not detected to 1 ng/g fat for hexaBDE, 0.001–2 ng/g fat for heptaBDE and 

not detected to 8 ng/g fat for octaBDE. NonaBDE concentrations were estimated to be >1 ng/g fat; 

decaBDE was estimated to range between not detected and 0.7 ng/g fat. In the late 1990s, breast adipose 

samples collected in northern California contained quantifiable amounts of 2,2’,4,4’-tetraBDE (BDE 47), 

2,2’,4,4’,5-pentaBDE (BDE 99), and 2,2’,4,4’,5,5’-hexaBDE (BDE 153) (She et al. 2000). Mean 
PBBs and PBDEs 448 

8. POTENTIAL FOR HUMAN EXPOSURE 

concentrations were 18 ng/g lipid weight for BDE 47, 4.9 ng/g lipid weight for BDE 99, and 2.2 ng/g 

lipid weight for BDE 153. Average total PBDEs levels (86 ng/g lipid) were the highest human levels 

reported to date. Recently, Petreas et al. (2003) expanded their investigation to include a diverse group of 

local women from the San Francisco Bay area sampled in the late 1990s. Their results confirmed earlier 

findings reported in She et al. (2000). Mean concentrations of BDE 47 in adipose tissues samples taken 

from California women were 28.9 ng/g lipid. In the adipose tissue of a 74-year-old Swedish male, the 

BDE 47 concentration was 8.8 ng/g lipid weight (Haglund et al. 1997). 

Adipose and liver tissue from two Swedish males were examined for several PBDEs (2,4,4’-triBDE 

[BDE 28], 2,2’,4,4’-tetraBDE [BDE 47], 2,2’,3,4,4’-pentaBDE [BDE 85], 2,2’,4,4’,5-pentaBDE 

[BDE 99], 2,2’,4,4’,6-pentaBDE [BDE 100], 2,2’,4,4’,5,5’-hexaBDE [BDE 153], and 2,2’,4,4’,5,6’-hexaBDE [BDE 154]) (Meironyté Guvenius and Norén 1999). The distribution of congener concentrations in 

the adipose and liver tissues for each individual were similar. BDE 47, BDE 99, and BDE 153 were the 

predominant congeners with adipose BDE 47 concentrations ranging from 2 to 2.4 ng/g lipid weight, 

BDE 99 concentrations of 1.6 ng/g lipid weight, BDE 100 concentrations of 0.1 ng/g lipid weight, and 

BDE 153 concentrations ranging from 1 to 1.3 ng/g lipid weight. The total PBDE concentration (i.e., the 

sum of the seven congeners) in adipose tissue was 5 ng/g lipid weight. Human liver and adipose tissues 

from one woman and four men autopsied in Sweden in 1994 were analyzed for PBDEs containing 3–

6 bromine atoms (Meironyté Guvenius and Norén 2001). PBDEs were found in all of the tissue samples. 

The sums of nine congeners (BDE 17, BDE 28, BDE 47, BDE 66, BDE 100, BDE 99, BDE 85, 

BDE 154, and BDE 153) were 5–18 and 4–8 ng/g lipids in liver and adipose tissue, respectively. The 

PBDE congeners BDE 47, BDE 99, and BDE 153 occurred at the highest levels and constituted 87–

96 and 84–94% of the total sum in liver and adipose tissue, respectively. Strandman et al. (1999) 

measured the concentration of BDE 47, BDE 99, and BDE 153 in adipose tissue samples from 

10 randomly selected individuals in Finland. Mean concentrations were 7.3 ng/g fat for BDE 47, 2.2 ng/g 

fat for BDE 99, and 2.3 ng/g fat for BDE 153. Levels of PBDEs were measured in adipose tissue samples 

from 13 individuals (3 women, 10 men) from Tarragona, Spain; the mean concentrations of BDE 47, 

BDE 99, and BDE 153 were 1.36, 0.42, and 1.83 ng/g lipid weight, respectively. The mean 

concentrations of PeBDE and HexaBDE were 0.93 and 1.83 ng/g lipid weight, respectively (Meneses et 

al. 1999). 

Human Milk. Recently, Schecter et al. (2003b) reported the first findings on levels of PBDEs congeners 

in human milk from individuals in the United States. Forty-seven individual milk samples were analyzed 

from nursing mothers, 20–41 years age, from a milk bank in Austin, Texas, and a community health clinic 
PBBs and PBDEs 449 

8. POTENTIAL FOR HUMAN EXPOSURE 

in Dallas, Texas, both in the year 2001. The median concentration of the sum of PBDE congeners was 

34.0 ng/g lipid. The predominant congener was BDE 47 (18.4 ng/g lipid); other congeners detected were 

2,2’,4-triBDE (BDE 17), 2,4,4’-triBDE (BDE 28), 2,3’,4,4’-tetraBDE (BDE 66), 2,2’,3,4,4’-pentaBDE 

(BDE 85), 2,2’,3,4,4’-pentaBDE (BDE 99), 2,2’,4,4’,6-pentaBDE (BDE 100), 2,2’,3,4,4’,5’-hexaBDE 

(BDE 138), 2,2’,4,4’,5,5’-hexaBDE (BDE 153), 2,2’,4,4’,5,6’-hexaBDE (BDE 154), and 

2,2’,3,4,4’,5’,6-heptaBDE (BDE 183) at median concentrations of 0.01, 1.2, 0.14, 0.41, 5.7, 2.9, 0.09, 2.0, 

0.22, and 0.07 ng/g lipd, respectively. DecaBDE was detected in 7 out of 47 samples with a maximum 

concentration of 8.24 ng/g lipid. PBDE levels in breast milk from this study were similar to levels found 

in U.S. blood and adipose tissue lipid from California and Indiana and are 10–100 times greater than 

human tissue levels in Europe (Schecter et al. 2003b). 

Norén and Meironyté (1998, 2000) examined the temporal trends of PBDE concentrations in pooled 

breast milk samples from mothers in Stockholm, Sweden. Between 1972 and 1997, the concentration of 

PBDEs in human breast milk increased, with a doubling rate of 5 years. In the 1997 sample, the 

concentration of PBDEs (sum of eight congeners) was 4 ng/g lipid, whereas the 1972 sample contained 

0.07 ng/g lipids (Meironyté et al. 1999). The authors suggest that the current exposure of humans to 

PBDEs may not be only diet; other exposure routes may result from the presence of PBDE in both work 

and home environments. PBDE levels were studied in breast milk obtained from mothers pregnant for 

the first time (n=39, ages 22–36 years old) from Uppsala County, Sweden (Darnerud et al. 1998). The 

mean value of total PBDEs (sum of eight congeners) was 4.4 ng/g fat; the major congener was BDE 47, 

contains ca. 55% of the total PBDEs. Recently, Lind et al. (2003) reported levels of PBDEs in human 

breast milk sampled from Uppsala County, Sweden. Total PBDEs, 2,2’,4,4’-tetraBDE (BDE 47), 

2,2’,4,4’,5-pentaBDE (BDE 99), and 2,2’,4,4’,6-pentaBDE (BDE 100) levels were 4.01, 2.35, 0.62, and 

0.38 ng/g lipid, respectively. In human breast milk from 25 German mothers, the levels of PBDEs ranged 

from 0.6 to 11 ng/g lipid (de Wit 2002). In 1992, the mean concentration of total PBDEs (sum of 

2,4,4-triBDE [BDE 28], -47, -99, -100, 2,2’,4,4’,5,5’-hexaBDE [BDE 153], and 2,2’,3,4,4’,5’,6-heptaBDE [BDE 183]) was 5.8 ng/g lipid weight for samples (n=6) from mothers from Ontario and Quebec, 

Canada (Ryan and Patry 2000). Combined samples from 1992 representing four regions of Canada and 

one representing all Canadian provinces had total PBDE concentrations ranging from 2.6 to 19 ng/g lipid 

weight; the highest concentrations were observed in the New Brunswick, Nova Scotia, and Prince Edward 

Island. Breast milk samples from Finland, collected between 1994 and 1998, had concentrations of total 

PBDEs (sum of BDE 28, BDE 47, BDE 99, and BDE 153) ranging from 0.88 to 5.9 ng/g lipid weight 

(Strandman et al. 2000). In Japan, breast milk samples had total PBDE concentrations (sum of BDE 28, 

BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) ranging from 0.66 to 2.8 ng/g lipid weight (Ohta et 
PBBs and PBDEs 450 

8. POTENTIAL FOR HUMAN EXPOSURE 

al. 2002a). Women who consumed fish had a somewhat higher concentration of total PBDEs (range, 1.4–

2.8 ng/g lipid weight) compared to women who ate less fish (range, 0.67–0.87 ng/g lipid weight). 

BDE 47 was the major congener in most of the samples; BDE 153 levels were analogous to BDE 47 

levels in some samples (Ohta et al. 2002a). 

Hydroxy- and Methoxy- Derivatives in Biota. Hydroxy- and methoxy- derivatives of PBDEs have been 

identified in biota. However, their orgin in the environment has not yet been explained. Anthropogenic 

sources of these compounds have not been found. Tetra- and pentabrominated methoxy (MeO) BDEs 

were found in herring, salmon, grey seal, ringed seal, and white-tailed sea eagle from the Baltic region 

(Asplund et al. 1999a; Haglund et al. 1997; Olsson et al. 2000) as well as beluga whale from Svalbard and 

pilot whale from the Faroe Islands (van Bavel et al. 2001). The concentrations of hydroxy- and methoxyderivatives were of the same order of magnitude as PBDEs present in the samples. Biogenic production 

via metabolism of PBDEs or natural production via biobromination have been suggested as the origin for 

these compounds. Naturally produced methoxy-tetrabrominated diphenyl ethers have been reported in 

tropical marine sponges (sp. Dysidea) as well as in green algae (sp. Cladophora) collected in Japan 

(Kierkegaard et al. 2004). Kierkegaard et al. (2004) found that the concentrations of 6-methoxy2,2’,4,4’-tetrabromodiphenyl etherin herring from five locations along the Swedish coast increased from 

south to north in the Baltic Sea. No correlation between the concentrations of BDE congeners and 

methoxy-brominated diphenyl ethers was observed, indicating sources other than PBDEs for these 

compounds. 

8.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE 

Polybrominated Biphenyls. PBBs are no longer produced or used in the United States. Thus, the general 

population exposure to PBBs will only be from historical releases. For people residing in the lower 

peninsula of Michigan, especially in the immediate vicinity of the PBB contaminated areas of this region, 

exposure to PBBs may still be occurring today. However, environmental levels have decreased since the 

1970s and current exposure, if any, will be at low levels. For other regions of the United States, the levels 

of exposure will either be very low or none. 

In the past, the general population may have been exposed to PBBs by inhaling contaminated air, 

ingesting contaminated water and food, and using consumer products containing PBBs. Other than in air 

in the vicinity of PBB production plants (see Section 8.1), no current or historical data exist that would 

indicate that PBBs might be present in ambient air. There are no current or historical data on the direct 
PBBs and PBDEs 451 

8. POTENTIAL FOR HUMAN EXPOSURE 

exposure of humans to PBBs from water. The general population may have been exposed to low levels of 

PBBs from the consumption of contaminated foods, but no estimate is available that quantitated this 

exposure. Historical monitoring and body burden data indicate that low level exposures to PBBs were 

limited to the population within the state of Michigan (see Section 8.4 and Table 8-5). The level of 

exposure to PBBs was slightly higher for the people residing in the lower peninsula of Michigan and 

highest among people residing in the immediate vicinity of the contaminated dairy farms, where people 

consumed contaminated meat, eggs, and dairy products (see Section 8.4 and Table 8-5). Consumer 

exposure in the past (plastics containing PBBs may not be in circulation anymore since PBB production 

ceased in the 1970s) from using PBB-containing plastics (e.g., typewriters, calculators, projector 

housings, and movie equipment cases) is expected to be very low since the PBBs were incorporated into 

the plastic and their mobilization could only have occurred under conditions such as combustion (Di 

Carlo et al. 1978). 

Workers involved in the historical production of PBBs, PBB-containing plastics, and PBB-containing 

plastic products could have been exposed to PBBs via inhalation of dust and vapor and/or dermal contact. 

Both workplace environmental monitoring and body burden monitoring data of workers (see Table 8-5) 

(Hesse and Powers 1978; Humphrey and Hayner 1975; Wolff et al. 1979b) indicated that workers in PBB 

industries were exposed to higher concentrations of PBBs than the general population. Although no 

evidence has been reported, workers in facilities that combusted or incinerated PBB-containing plastics 

might have been exposed to higher levels of PBBs. 

Polybrominated Diphenyl Ethers. Body burden data indicate that there are low-level exposures to 

PBDEs for the general population. However, the current understanding of exactly how low levels of 

certain PBDE isomers/congeners (e.g., 2,2’,4,4’-tetraBDE [BDE 47]; 2,2’,4,4’,5-pentaBDE [BDE 99]; 

2,2’,4,4’,6-pentaBDE [BDE 100]) came to be present in human tissues is insufficient to reach definitive 

conclusions. Humans appear to be exposed to lower brominated BDEs by ingestion of contaminated 

foods (e.g., fish) and possibly inhalation of ambient or contaminated air. Dermal exposure to PBDEs 

could occur by contact with products containing PBDEs such as textiles or polymers. Inhalation exposure 

could occur from outgassing of PBDEs from electrical appliances and furniture into indoor atmospheres. 

However, little information is known about the potential exposure from these routes. 

In the United States, exposure is evident by the levels of lower brominated PBDEs found in tissues from 

individuals (see Section 8.4.4). For example, breast adipose samples collected in northern California in 

the late 1990s contained quantifiable amounts of 2,2’,4,4’-tetraBDE (BDE 47), 2,2’,4,4’,5-pentaBDE 
PBBs and PBDEs 452 

8. POTENTIAL FOR HUMAN EXPOSURE 

(BDE 99), and 2,2’,4,4’,5,5’-hexaBDE (BDE 153) (She et al. 2000). Mean concentrations were 18 ng/g 

lipid weight for BDE 47, 4.9 ng/g lipid weight for BDE 99, and 2.2 ng/g lipid weight for BDE 153. In 

studies of the general populations of other countries, it has also been shown that exposure to lower 

brominated PBDE congeners by the general population is widespread (see Section 8.4.4; Haglund et al. 

1997; Meneses et al. 1999). In general, levels of decaBDE in human tissues and body fluids are 

negligible which indicates that direct exposure to decaBDE congener appears to be low. However, the 

level current information is insufficient to draw conclusions on whether decaBDE degrades in the 

environment and if possible degradation products of decaBDE may be important sources of human 

exposure. 

The concentration of total PBDEs in outdoor air ranges from 2 (rural) to 77 (urban) pg/m3

 in the United 

States (Dodder et al. 2000a). Typically, lower brominated congeners (e.g., 2,2’,4,4’-tetraBDE [BDE 47]) 

are predominant which indicates low levels of exposure to these congeners by the general population. 

Levels of higher brominated congeners (e.g., decaBDE) tend to be below the limit of quantitation. In 

Sweden, indoor air concentrations of PBDEs in lecture halls, computerized indoor environments, and 

rooms with electronic devices (e.g., televisions) have low levels of PBDEs (Lindström 1999). Point 

sources may result in increased concentrations of PBDEs in indoor air. For example, octaBDE has been 

found in indoor areas that contain electronic products containing PBDEs (e.g., televisions and computers) 

(Bergman et al. 1997). The release of PBDEs from polymers is dependant on the migration ability of the 

PBDE molecule through the polymer matrix to the polymer surface where emission is possible (Danish 

EPA 1999). Because PBDEs are large molecules, migration is expected to occur slowly. No 

experimental studies were located on the emission rate of PBDEs from plastics. Based on worst case 

emission factors, the estimated emission of deca-, octa-, and pentaBDEs from plastics are 0.038, 0.054, 

and 0.39% per year, respectively (Danish EPA 1999). 

Harrad et al. (2004) found a significant possitive correlation between PBDE concentrations in indoor air 

and both the number of electrical appliances and the number of chairs containing polyurethane foam. 

Concentrations of tetra- and pentabrominated concengers (BDE 47, 99, and 100) in indoor air were 

always higher than those detected in outdoor air. On average, indoor air concentrations were 150, 120, 

and 140 times higher than outdoor air for BDE 47 (1,700 pg/m3

), 99 (852 pg/m3), and 100 (217 pg/m3), 

respectively. In this study, the median lower bound estimates of inhalation exposure (i.e., where a 

congener is below the detection limit, the concentration is assumed to be zero) for BDE 47 

(2,2’,4,4’-tetraBDE), 99 (2,2’,4,4’,5-pentaBDE), 100 (2,2’,4,4’,6-pentaBDE), 153 (2,2’,4,4’,5,5’-hexa-
PBBs and PBDEs 453 

8. POTENTIAL FOR HUMAN EXPOSURE 

BDE), 154 (2,2’,4,4’,5,6’-hexaBDE), and total PBDEs were 4.5, 1.2, 0.41, 0.016, 0.040, and 6.9 ng/day, 

respectively (Harrad et al. 2004). 

Consumption of food is expected to be the major route of exposure in humans (Lindström 1999). 

Consumption of fish has been associated with elevated levels of PBDEs in tissues from the Swedish 

population (Bergman et al. 1999). In Sweden, fish consumption is about 30 g/day; this translates to an 

estimated 0.1 μg of pentaBDE and 0.3 μg of total PBDEs from fish that is ingested by humans daily 

(WHO 1994a). The fish of greatest concern to humans are bottom feeders like carp and catfish. Harrad 

et al. (2004) estimated the daily dietary intakes of PBDEs in omnivorous and vegetarian diet samples 

from the United Kingdom. In this study, the median lower bound estimates of dietary exposure (i.e., 

where a congener is below the detection limit, the concentration is assumed to be zero) for BDE 47 

(2,2’,4,4’-tetraBDE), 99 (2,2’,4,4’,5-pentaBDE), 100 (2,2’,4,4’,6-pentaBDE), 153 (2,2’,4,4’,5,5’-hexaBDE), 154 (2,2’,4,4’,5,6’-hexaBDE), and total PBDEs were 46.4, 42.6, 0, 0, 0, and 90.5 ng/day, 

respectively (Harrad et al. 2004). Like PCBs, there may be a higher risk of exposure to PBDEs in Native 

Americans who reside in the Arctic region and consume whale and seal blubber (Jaret 2000). 

Workers involved in the production and manufacture of PBDE-containing plastics and plastic products 

are exposed to PBDEs. Body burden data indicate higher levels for workers exposed to PBDEs than for 

the general population. Occupational exposure to PBDEs also occurs in workers at plants that dismantle 

electronic equipment, computer monitor repair technicians, and automobile drivers, as well as other 

professions (Lindström 1999). Occupational exposure occurs primarily by inhalation. Inhalation of 

vapor phase PBDEs is expected to be low due to the low vapor pressures of PBDEs (see Table 6-6); 

however, the inhalation of particulate phase PBDEs is possible during plastic reprocessing where grinding 

or shredding of polymers with PBDEs occurs. Occupational exposure may also likely involve oral 

exposure to particulate PBDEs as a result of hand-to-mouth activity. 

Air samples were taken from an electronics dismantling plant, an office with computers, and outdoors and 

then analyzed for PBDEs (Sjödin et al. 1999a, 2001a). The electronics dismantling plant had the highest 

concentrations of PBDEs, with mean concentrations of 2.5 pmol/m3

 (1.25 ng/m3) for 2,2’,4,4’-tetraBDE 

(BDE 47), 4.6 pmol/m3

 (2.6 ng/m3) for 2,2’,4,4’,5-pentaBDE (BDE 99), 6.1 pmol/m3 (3.93 ng/m3) for 

2,2’,4,4’,5,5’-hexaBDE (BDE 153), 26 pmol/m3

 (18.8 ng/m3) for 2,2’,3,4,4’,5’,6-heptaBDE (BDE 183), 

and 38 pmol/m3

 (36.5 ng/m3) for decaBDE (BDE 209) (Sjödin et al. 1999a, 2001a). Air samples were 

found to be 4–10 times higher in PBDE concentrations near a plastic shredder when compared to other 

locations in the plant (range, 0.42–200 ng/m3

). Concentrations of PBDEs in the office (range, <0.002–
PBBs and PBDEs 454 

8. POTENTIAL FOR HUMAN EXPOSURE 

0.09 ng/m3) were 400–4,000 times lower than in the plant, and PBDEs were not detected in outside air 

(Sjödin et al. 1999a, 2001a). 

2,2’4,4’-TetraBDE (BDE 47), 2,2’,4,4’,5,5’-hexaBDE (BDE 153), 2,2’,4,4’,5,6’-hexaBDE (BDE 154), 

2,2’,3,4,4’,5’,6-heptaBDE (BDE 183), and decaBDE (BDE 209) were measurable in blood plasma from 

three groups of workers (i.e., workers at a computer disassembly plant, workers in a computerized office, 

and a control group) (Sjödin et al. 1999b). The median concentrations (sum of five congeners) were 

highest for the computer disassembly plant workers (26 ng/g lipid weight); the office workers had a 

median concentration of 4.1 ng/g lipid weight and the control group had a median concentration of 

3.3 ng/g lipid weight. The congener patterns for the control group and office workers were similar, with 

BDE 47 having the highest levels. For the computer disassembly plant workers, the median 

concentrations of BDE 183, BDE 153, BDE 154, BDE 47, and BDE 209 were 7.8, 4.5, 1.2, 2.9, and 

4.8 ng/g lipid weight, respectively. Blood serum samples from 19 full-time computer technicians were 

analyzed (Hagmar et al. 2000). The serum concentrations of BDE 153, BDE 183, and BDE 209 in these 

samples were found to be approximately 5 times higher than the control and office workers in the Sjödin 

et al. (1999b) study. The median concentration for total PBDEs (for the sum of five congeners) was 

10.6 pmol/g (7.0 ng/g) lipid weight. The highest concentrations were for BDE 153. Two octaBDE 

congeners and one nonaBDE congener were also detected. 

8.6 EXPOSURES OF CHILDREN 

This section focuses on exposures from conception to maturity at 18 years in humans. Differences from 

adults in susceptibility to hazardous substances are discussed in Section 5.7 Children’s Susceptibility. 

Children are not small adults. A child’s exposure may differ from an adult’s exposure in many ways. 

Children drink more fluids, eat more food, breathe more air per kilogram of body weight, and have a 

larger skin surface in proportion to their body volume. A child’s diet often differs from that of adults. 

The developing human’s source of nutrition changes with age: from placental nourishment to breast milk 

or formula to the diet of older children who eat more of certain types of foods than adults. A child’s 

behavior and lifestyle also influence exposure. Children crawl on the floor, put things in their mouths, 

sometimes eat inappropriate things (such as dirt or paint chips), and spend more time outdoors. Children 

also are closer to the ground, and they do not use the judgment of adults to avoid hazards (NRC 1993). 
PBBs and PBDEs 455 

8. POTENTIAL FOR HUMAN EXPOSURE 

Polybrominated Biphenyls. Infants who consume breast milk may have had a higher exposure to PBBs 

than children who drink formula milk, especially children exposed during the Michigan episode (see 

Section 8.4.4). No additional information was found in the literature about the exposure of children to 

PBBs (WHO 1994b). 

Polybrominated Diphenyl Ethers. Infants who consume breast milk may have a higher exposure to 

lower brominated BDEs than children who drink formula milk (see Section 8.4.4). Exposure of neonates 

is evident due to the presence of lower brominated BDEs (e.g., 2,2’,4,4’-tetraBDE [BDE 47]) in breast 

milk and placental tissue. The PBDEs detected in breast milk are the tri- to hexaBDEs, but not the heptato decaBDEs (LaKind and Berlin 2000), which are the same congeners found in bioaccumulation studies 

with fish and other mammals. Schecter et al. (2003b) reported the first findings on levels of PBDEs 

congeners in human milk from individuals in the United States. The median concentration of the sum of 

PBDE congeners was 34.0 ng/g lipid with BDE 47 (18.4 ng/g lipid) as the predominant congener. 

DecaBDE was detected in 7 out of 47 samples with a maximum concentration of 8.24 ng/g lipid. The 

levels of PBDEs in breast milk from this study were 10–100 times greater than human tissue levels in 

Europe (Schecter et al. 2003b). Levels of lower brominated BDEs in the breast milk of Swedish women 

shows an exponentially increasing trend in exposure since the 1970s, with concentrations of lower 

brominated BDEs in breast milk doubling every fifth year (Lindström 1999). The concentrations of lower 

brominated BDEs in breast milk are increasing exponentially from about 300 pg/g lipid in 1976 to about 

4,000 pg/g lipid in 1997 in Swedish women (Norön and Meironyté 2000). The sum of four PBDE 

congeners (2,4,4’-triBDE [BDE 28], BDE 47, 2,2’,4,4’,5-pentaBDE [BDE 99], and decaBDE [BDE 209]) 

was between 0.88 and 5.89 ng/g lipid in breast milk and between 1.00 and 4.40 ng/g lipid in placental 

tissue of 11 Finnish women (Strandman et al. 2000). The four highest concentrations of total PBDEs 

were found in nulliparous women. In 2001, Mazdai et al. (2003) determined the concentration of six 

PBDE congeners (BDE 47, BDE 99, 2,2’,4,4’,6-pentaBDE [BDE 100], BDE 153, 2,2’,4,4’,5,6’-hexaBDE 

[BDE 154], and 2,2’,3,4,4’,5’,6-heptaBDE [BDE 183]) and total PBDEs in maternal and fetal blood 

samples taken from subjects in the United States. Median levels of total PBDE (sum of six congeners) 

were 39 and 37 ng/g lipid for fetal and maternal serum, respectively. BDE 47 was the predominant 

congener reported at median concentrations of 25 and 28 ng/g lipid for fetal and maternal serum samples, 

respectively. Concentrations of PBDEs in fetal and maternal serum were comparable. Thus, the authors 

concluded that measurement of maternal serum PBDE levels could be used to determine fetal exposure to 

PBDEs. Concentrations of PBDEs were determined in children from Norway (Thomsen et al. 2002b). 

Samples were collected between the period of 1975–2002. Children ages 0–4 years had congener levels 

of BDE 28, BDE 47, BDE 100, BDE 99, 2,2’,4,4’,5,5’-hexaBDE (BDE 153), and BDE 154 at 0.26, 6.2, 
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8. POTENTIAL FOR HUMAN EXPOSURE 

1.7, 1.6, 1.5, and 0.45 ng/g lipid, respectively; while children 4–14 years old had levels of BDE 28, 

BDE 47, BDE 100, BDE 99, BDE 153, and BDE 154 at 0.20, 2.0, 0.66, 0.37, 0.86, and 0.39 ng/g lipid, 

respectively. No additional information was found in the literature about the exposure of children to 

PBDEs (WHO 1994a). 

8.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES 

Polybrominated Biphenyls. The production of PBBs ceased in 1979, and the usable life of the plastics 

containing PBBs has expired. Therefore, these plastics are probably no longer in circulation. At the 

present time and in the near future, populations potentially exposed to low levels of PBBs are those living 

near hazardous waste sites in which the PBB-containing plastics have been disposed and the residents in 

and around the contaminated farms in Michigan. The lifetime of PBBs in soil is on the order of years 

(Jacobs et al. 1978), and the levels of PBBs in fish caught in contaminated waters have declined slowly 

(Hesse and Powers 1978). Therefore, concentrations of residual PBBs in soil and streams in the vicinity 

of PBB-containing hazardous waste sites, PBB production facilities, and contaminated farm areas are 

expected to remain above background levels for many years. The sources of potential exposure to PBBs 

for residents in these areas are consumption of contaminated meat and dairy products obtained from herds 

grazing over contaminated soil and consumption of fish from nearby contaminated streams. PBB 

contamination has triggered the issuance of one human health advisory in the state of Michigan. As of 

September 30, 1993, recreational and subsistence fishermen who consume appreciably higher amounts of 

fish caught in the Pine River downstream from St. Louis in Gratiot and Midland Counties (RTI 1993) 

may be exposed to above-average levels of PBBs associated with dietary intake (EPA 1993). The body 

burden for PBBs in residents of contaminated areas has been higher than in the general population 

(Brilliant et al. 1978; Cordle et al. 1978; Eyster et al. 1983; Humphrey and Hayner 1975; Kimbrough 

1987; Lambert et al. 1990; Landrigan et al. 1979; Wolff et al. 1982). Therefore, babies breast fed by 

exposed mothers in the contaminated areas may also be at higher risk (Jacobson et al. 1989). 

Polybrominated Diphenyl Ethers. Subsistence fishermen who consume PBDE-contaminated fish and 

Native Americans who reside in the Arctic region and consume whale and seal blubber may have a higher 

risk of exposure to lower brominated PBDEs (WHO 1994a). Other populations with high exposure levels 

to PBDEs involve occupational exposures (see Section 8.6). No other information was located that 

identified specific populations with higher exposure levels to PBDEs. 
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8.8 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the 

Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of PBBs and PBDEs are available. Where adequate 

information is not available, ATSDR, in conjunction with NTP, is required to assure the initiation of a 

program of research designed to determine the health effects (and techniques for developing methods to 

determine such health effects) of PBBs and PBDEs. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will be 

evaluated and prioritized, and a substance-specific research agenda will be proposed. 

8.8.1 Identification of Data Needs 

Physical and Chemical Properties. 

Polybrominated Biphenyls. Many of the relevant physical and chemical properties of the PBBs are not 

available (see Table 6-3). More data on the physical and chemical properties of hexabromobiphenyl are 

available relative to octabromo- and decabromobiphenyl. Even in the case of hexabromobiphenyl, not all 

relevant data are available, and the quality of data is questionable because the properties of FireMaster 

BP-6 have been reported as the properties of hexabromobiphenyls. More importantly, very limited data 

are available on the physical and chemical properties for the individual congeners of hexabromo-, octabromo-, and decabromobiphenyl. The absence of such important data as Koc, vapor pressure, and Henry’s 

law constant, is a major impediment in the prediction of the environmental fate and transport of PBBs. 

Polybrominated Diphenyl Ethers. Many of the relevant physical and chemical properties of the PBDEs 

are available (see Table 6-4). Very limited data are available on the physical and chemical properties for 

the individual congeners (Braekvelt et al. 2003; Tittlemier et al. 2002). Important data, such as Koc, vapor 

pressure, and Henry’s law constant, are necessary for the prediction of the environmental fate and 

transport of PBDEs. 
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8. POTENTIAL FOR HUMAN EXPOSURE 

Production, Import/Export, Use, Release, and Disposal. 

Polybrominated Biphenyls. The production of all PBBs in the United States stopped in 1979 (IARC 

1986). Data on the past production, import/export, and use of PBBs are available (Neufeld et al. 1977). 

In the past, PBB-containing plastic was used in consumer products, but the useful life of these products 

may have ended (Di Carlo et al. 1978; Neufeld et al. 1977), and these products are probably no longer in 

circulation. In the workplace, the environmental media contaminated by PBBs were air, water, and soil 

(DeCarlo 1979). Outside of the workplace, soil is expected to be the medium with significant 

contamination due to disposal of solid waste from production plants and disposal of PBB-containing 

plastics in landfills (Neufeld et al. 1977). Although it is known that PBB-containing plastics may have 

been disposed in landfills (Di Carlo et al. 1978), the amount that may have been incinerated is not known. 

No data were located from studies that determined the efficiency of incineration as a method of disposal 

of PBBs present in the neat form in industrial wastes or in plastics. Environmental regulations regarding 

the manufacture and disposal of PBBs have been established (EPA 1988a). According to the Emergency 

Planning and Community Right-to-Know Act of 1986, 42 U.S.C. Section 11023, industries are required 

to submit chemical release and off-site transfer information to the EPA. The Toxics Release Inventory 

(TRI), which contains this information for 1999, became available in 2002. This database will be updated 

yearly and should provide a list of industrial production facilities and emissions. 

Polybrominated Diphenyl Ethers. Production and use data are available for PBDE commercial mixtures 

(BSEF 2003). PBDEs are used as additive flame retardants in thermoplastics at levels ranging from 5 to 

30% by weight (EU 2001). The commercial pentaBDE product is used predominantly (95–98%) for 

flame-retardant purposes as an additive in consumer products manufactured by the furniture industry 

(ENVIRON 2003a). The commercial octaBDE is used by the plastics industry as an additive flame 

retardant for manufactured products. It is used almost exclusively as a flame retardant for acrylonitrilebutadiene-styrene (ABS) terpolymers used in computer casings and monitors (ENVIRON 2003b). The 

commercial decaBDE product is an additive flame retardant used in a variety of polymer applications. 

Industry information indicates that decaBDE is used at loadings of 10–15% weight in polymers and is 

always used in conjunction with antimony trioxide (EU 2002). The Great Lakes Chemical Corporation 

recently announced that it is voluntarily phasing out production of pentaDBEs and octaDBEs by the end 

of 2004 (Tullo 2003). In the United States, waste disposal of PBDE-containing consumer products is 

described as transfers to disposal (landfill), recycling, energy recovery (incineration), or publicly owned 

treatment works (POTWs) (Darnerud et al. 2001). More information for PBDEs is needed on 

import/export, release, and disposal. In particular, the mechanism by which PBDEs are leaving the 

products in which they are used and entering the environment is not understood. Possibilities include 
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8. POTENTIAL FOR HUMAN EXPOSURE 

disinigradtion of plastic products in particulates contaminated with PBDEs or volatilization of PBDEs 

from the plastic itself (Hites 2004). In the United States, waste disposal of PBDE-containing consumer 

products is described as transfers to disposal (landfills), recycling, energy recovery (incineration), or 

POTWs (Darnerud et al. 2001). Soil should be a medium of significant contamination due to disposal of 

PBDE-containing plastics in landfills. The commercial decaBDE and octaBDE products are used in hard 

dense plastics from which migration would be very difficult (BFRIP 2002). Although it is known that 

PBDE-containing plastics are disposed in landfills, information on the recycling of PBDE-containing 

plastics and the amounts of PBDE-containing plastics that are incinerated is not known. Additional 

information of levels of PBDEs and PBDD/PBDF from incineration of PBDE-containing plastics would 

be useful in determining exposure to the general population. 

Environmental Fate. 

Polybrominated Biphenyls. Information regarding the environmental fate of PBBs in air was not located 

in the literature. The data about the fate of PBBs in air are important for the prediction of transport 

characteristics of these compounds in air. Photolysis of the PBBs will produce debrominated products in 

proton-donating organic solvents (Ruzo and Zabik 1975; Ruzo et al. 1976), but there is less certainty 

about the importance of photolysis of PBBs in water (Norris et al. 1973; Ruzo et al. 1976). PBBs will 

partition from the aquatic phase to sediment and suspended solids in water (Hesse and Powers 1978). 

PBBs will bioconcentrate in aquatic organisms, but the BCF may decrease as the bromine substitution 

exceeds six (Gobas et al. 1989; Opperhuizen et al. 1985; Zitko 1979; Zitko and Hutzinger 1976). 

However, the difference in the reported BCF values for hexabromobiphenyl among different investigators 

is vast (Gobas et al. 1989; Hesse and Powers 1978; Opperhuizen et al. 1985; Veith et al. 1979). PBBs 

will remain strongly sorbed to soil (Filonow et al. 1976; Griffin and Chou 1981a, 1981b) and will persist 

in soil because of the lack of suitable degradation pathways (Jacobs et al. 1978). The translocation of 

PBBs from soil to upper parts in plants was not observed, and the transfer of PBBs from soil to carrot 

roots was found to be minor (Jacobs et al. 1976, 1978). A recent article by de Boer et al. (1998) found 

PBBs in deep ocean marine mammals, which suggests that PBBs may be transported globally. More 

monitoring data for PBBs in the environment are needed to verify the possible global transport of PBBs. 

Since the toxicity and the environmental fate of PBBs depends on specific PBBs congeners, development 

of more data regarding congener-specific fate and transport of PBBs in the environment are needed. 

Polybrominated Diphenyl Ethers. Based on limited data, photolysis appears to be the dominant 

transformation process for some PBDEs (e.g., decaBDE) (Hua et al. 2003). PBDEs absorb light in the 

environmental spectrum. Hua et al. (2003) found that decaBDE and the commercial octaBDE absorbed 
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8. POTENTIAL FOR HUMAN EXPOSURE 

light up to 325 nm, which indicates that these compounds may be susceptible to photodegradation at 

environmental wavelengths. However, the importance of photochemical transformation reactions in the 

environment cannot be determined due to lack of quantitative rate information (EU 2002, 2003). Based 

on a very limited number of studies, biodegradation does not appear to be significant for PBDEs 

commercial mixtures (EU 2002, 2003). Limited studies have been done on biodegradation of PBDEs in 

the environment under both aerobic and anaerobic conditions, especially studies investigating dehalogenation mechanisms (EU 2002, 2003). More studies are needed to determine conclusively if 

commercial PBDE mixtures, such as decaBDE, are degraded to lower brominated congeners (e.g., 

2,2’4,4’-BDE [BDE 47]), which appear to bioaccumulate in fish, animals, and humans (see Sections 8.4). 

Since the toxicity and the environmental fate of PBDEs depend on specific PBDEs congeners, 

development of more data regarding congener-specific fate and transport of PBDEs in the environment 

are needed. 

Bioavailability from Environmental Media. 

Polybrominated Biphenyls. Available information regarding the rate of absorption of PBBs following 

inhalation, oral, or dermal contact is discussed in the Toxicokinetics Section (Section 5.4). Although no 

data on the bioavailability of PBBs from inhalation of contaminated air, or ingestion of or dermal contact 

with water, or inhalation of or dermal contact with soil are available, the bioavailabilities from these 

routes of exposure are expected to be far less than 100% because these compounds strongly sorb to 

particulate matter and soil. The estimated bioavailability of higher brominated biphenyls is expected to 

be even lower than the less brominated biphenyls due to stronger sorption characteristics of the former 

compounds. The estimated bioavailability of PBBs by farm animals from ingestion of contaminated soil 

was 56–65% (Fries 1985a). Also, studies on many persistent halogenated aromatic compounds clearly 

show that they become progessively less bioavailable with time (Alexander 2000). Often, three-fourths or 

more of the concentration of such compounds is not bioavailable. Information on the possibility of the 

very low bioavailability of PBBs is needed. 

Polybrominated Diphenyl Ethers. The absorption and distribution of PBDEs as a result of inhalation, 

ingestion, and dermal exposure are discussed in Sections 5.3.1, 5.3.2, and 5.3.3. Studies that describe the 

bioavailability of PBDEs commercial mixtures and congeners from ambient air, surface water, and 

groundwater, or soil do not exist. Essentially no good data exist on the adsorption of PBDEs commercial 

mixtures and congeners with the exception of decaBDE (BFRIP 2002). Studies on many persistent 

halogenated aromatic compounds clearly show that they become progessively less bioavailable with time 

(Alexander 2000). Often, three-fourths or more of the concentration of such compounds is not 
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8. POTENTIAL FOR HUMAN EXPOSURE 

bioavailable. Information on the possibility of the very low bioavailability of PBDEs is needed. Studies 

determining the effect of particle size and organic matter content on the bioavailability of PBDEs from 

soil and the role of microparticle-sorbed PBDEs on the bioavailability of PBDEs from drinking water are 

needed. Such studies would be useful in assessing the health effects of PBDEs on people living near 

hazardous waste sites. 

Food Chain Bioaccumulation. 

Polybrominated Biphenyls. PBBs do not readily translocate from soil to plants via root uptake (Jacobs et 

al. 1976, 1978). Therefore, PBBs may not bioconcentrate in plants. However, plant uptake data are 

limited, and it will be helpful to develop additional plant uptake data. Brominated biphenyls with 

bromine substitution 6 or less will bioconcentrate in aquatic organisms (Gobas et al. 1989; Norris et al. 

1973; Opperhuizen et al. 1985; Zitko 1979; Zitko and Hutzinger 1976). PBBs are preferentially stored in 

the adipose tissue of animals (Kimbrough 1987). Although PBBs have been detected in fish-eating birds 

and predatory animals from the consumption of contaminated food (Heinz et al. 1983, 1985; Hesse and 

Powers 1978), no systematic study was located that analyzed the biomagnification potential in predators 

resulting from consumption of contaminated food. 

Polybrominated Diphenyl Ethers. An abundance of monitoring data illustrate the uptake of lower 

brominated diphenyl ethers by aquatic organisms, which results in bioconcentration (see Section 8.4.4; 

Hardy 2002b). Congener components of the pentaBDE commercial product tend to bioconcentrate to 

different extents. DecaBDE and octaBDE commercial products appear to not bioconcentrate, 

bioaccumulate, or biomagnify (Hardy 2002b). The limited existing data indicate that lower brominated 

BDE congeners (e.g., pentaBDE commercial mixtures) bioaccumulate in aquatic and terrestrial food 

chains and biomagnify in predators due to consumption of contaminated prey. Bioaccumulation of 

PBDEs in the aquatic food web is inversely related to the degree of bromination (Burreau et al. 2000; 

Jansson et al. 1993). More information on bioaccumulation and biomagnification of PBDE and its 

congeners is needed in assessing human health risks. 

Exposure Levels in Environmental Media. 

Polybrominated Biphenyls. Only limited data on the levels of PBBs in ambient air are available 

(DeCarlo 1979). Data are available on the levels of PBBs in effluent water from manufacturing plants, in 

river water, stream sediment, and soil in the vicinity of the plants, in sludge of a waste treatment plant, 

and in groundwater of a landfill site (Hesse and Powers 1978; Shah 1978). No data on the level of PBBs 
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8. POTENTIAL FOR HUMAN EXPOSURE 

in drinking water from the contaminated sites were located. No estimate on the human intake of PBBs 

from any of the various environmental media was located in the literature. 

Reliable monitoring data for the levels of PBBs in contaminated media at hazardous waste sites are 

needed so that the information obtained on levels of PBBs in the environment can be used in combination 

with the known body burden of PBBs to assess the potential risk of adverse health effects in populations 

living in the vicinity of hazardous waste sites. 

Polybrominated Diphenyl Ethers. Information on the relative importance of different routes of exposure 

to PBDEs is limited especially in the United States. Atmospheric concentrations of PBDEs tend to be 

dominated by lower brominated congeners, e.g., 2,2’,4,4’-pentaBDE (BDE 47) (Dodder et al. 2000a; 

Strandberg et al. 2001). More monitoring data on the concentrations of total PBDEs and PBDE 

congeners in air in remote, rural, and urban areas, as well as areas near hazardous waste sites and 

incinerators are needed. Due to the hydrophobic nature of PBDEs, this class of compounds has not been 

detected in water to any significant extent. BDE 47 was detected at low concentrations in Lake Ontario 

surface waters (Luckey et al. 2001). Although levels are predicted to be low, monitoring data on PBDE 

concentrations in finished drinking water nationwide would be helpful. No information was located on 

the ambient environmental concentrations of PBDEs in soils in the United States or other parts of the 

world. Hale et al. (2002) reported the concentrations of PBDEs in soil samples collected in the vicinity of 

a polyurethane foam manufacturing facility, which are higher than expected in rural and potentially urban 

areas. Sediment concentrations of PBDEs tend to be dominated by higher brominated congeners (e.g., 

decaBDE or BDE 209) (deWit 2002; Dodder et al. 2002; Hale et al. 2001b, 2002); temporal trends 

suggest that concentrations of PBDEs in sediments are increasing. Information about the concentrations 

of PBDEs in food stuffs is very limited (Bocio et al. 2003; Huwe et al. 2002a; Ohta et al. 2002a), 

especially in the United States. Data on the concentrations of PBDEs in foods, collected using a marketbasket approach, are needed to determine concentrations of PBDEs in foods consumed by the general 

population. Data on the PBDE concentrations in foods grown in contaminated areas, particularly in the 

vicinity of hazardous waste sites, are also needed. Data on congener-specific PBDE analysis of food, 

especially plant products, would be useful. Monitoring data indicated that the levels of PBDEs are 

increasing in aquatic organisms with higher concentrations near point sources (Alaee et al. 1999; Dodder 

et al. 2000a; Johnson and Olson 2001; Loganathan et al. 1995; Luross et al. 2000). Additional monitoring 

data on environmental levels of PBDEs would to useful to determine the extent of contamination in 

environmental media, and also the mechanisms of human exposure to this class of chemicals. 
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8. POTENTIAL FOR HUMAN EXPOSURE 

Exposure Levels in Humans. 

Polybrominated Biphenyls. Body burden data indicate that low-level exposures to PBBs have occurred 

for people in the state of Michigan. No recent information about average daily intake of PBBs was 

located. The levels of PBBs in human tissue and body fluids, such as blood, serum, adipose tissue, breast 

milk, feces, cord blood, biliary fluid, and placenta, of people in the state of Michigan have been 

extensively studied (Brilliant et al. 1978; Cordle et al. 1978; Eyster et al. 1983; Humphrey and Hayner 

1975; Lambert et al. 1990; Landrigan et al. 1979; Wolff et al. 1982). However, no recent data are 

available. Data on the levels of PBBs in tissues and body fluids of residents in the vicinity of sites of 

industrial discharge of PBB wastes were not located. Updated information would be useful to understand 

current exposure levels of people in the state of Michigan to PBBs. This information is necessary for 

assessing the need to conduct health studies on these populations. 

Polybrominated Diphenyl Ethers. Body-burden data indicate that there are low-level exposures to lower 

brominated PBDEs for the general population. The absence of decaBDE (BDE 209) in the ambient 

population is likely the result of analytical bias since most studies of the ambient population did not 

include BDE 209 as one of the analytes of interest. Thus, future studies that include the analysis of 

BDE 209 in body tissues and fluids would be useful (Hites 2004). Information about the average daily 

intake of PBDEs is limited to populations living in Sweden (Bergman et al. 1999; Lindström 1999; WHO 

1994a). PBDE levels are reported in the current literature for blood, breast milk, and adipose tissue of the 

general population and occupationally exposed individuals (WHO 1994a). Limited information on the 

levels of PBDEs in body tissues and fluids from individuals living in the United States have been located 

(Mazdai et al. 2003; Patterson et al. 2000; Petreas et al. 2002, 2003; Schecter et al. 2003b; Sjödin et al. 

2001). These studies indicate that levels of lower brominated BDEs in body fluids are a factor of 10–

100 higher for individuals living in the United States compared to individuals living in other regions of 

the world (e.g., Europe). Limited surveys have ever been conducted in the United States to evaluate the 

trend of PBDE concentrations in human tissues over the years. It would be helpful to develop a database 

of information on congener-specific PBDE levels in tissues of exposed and control cases for studying 

clinical and epidemiological outcomes. In particular, a comprehensive study that monitors congenerspecific concentrations in fish species and relates them directly to congener levels in human tissue would 

be extremely useful. Monitoring data indicate 2,2’,4,4’-tetraBDE (BDE 47) and 2,2’,4,4’,5,5’-hexaBDE 

(BDE 153) are relatively higher in human samples than in the commercial pentaBDE product and that 

2,2’,4,4’,5-pentaBDE (BDE 99) is relatively lower. The cause of this difference is not known. It may be 

due to its relatively higher vapor pressure of BDE 99, or it may be due to the selective environmental 

elimination of BDE 99, a process that has been observed in some biota (Hites 2004). Additional studies 
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8. POTENTIAL FOR HUMAN EXPOSURE 

that examine these trends would be useful. Additional data regarding the concentrations of PBDEs in 

body fluids or tissues of people who reside near hazardous waste sites are needed. This information is 

necessary for assessing the need to conduct health studies on these populations. 

Exposures of Children. 

Polybrominated Biphenyls. Children may be exposed to PBBs by a variety of exposure pathways. 

Levels will be highest for children living in the vicinity of the area affected by the Michigan 

contamination episode. The most important pathway appears to be consumption of contaminated 

mother’s milk (see Section 8.4.4). More data are needed on the levels of PBB exposure in nursing 

women from consumption of fish and from those of the general population. Exposure and body burden 

studies related to consumption of fish in the U.S. population are needed to determine exposure levels, 

particularly in children of recreational and subsistence fishers. Information related to the exposure of 

children living near hazardous waste sites is also needed. In particular, information is needed that is 

related to the potential for children to be exposed to PBBs bound to soil and dust particles through pica or 

unintentional hand-to-mouth activity within homes located in these areas. Quantitative information 

regarding the bioavailability and amount of PBBs that children are exposed to through contact with 

contaminated soils are unavailable. Therefore, any information concerning this subject would be useful in 

evaluating children’s exposure. Additional information on weight-adjusted intakes would be helpful for 

determining the health risks for young children. Infants and young children consume a greater amount of 

food per kilogram of body weight and, therefore, may have a proportionately greater exposure to PBBs 

than adults. 

Polybrominated Diphenyl Ethers. Children may be exposed to PBDEs by a variety of exposure 

pathways. The most important pathway appears to be consumption of contaminated foods, particularly 

fish (Bergman et al. 1999; Lindström 1999; WHO 1994a). Children can also be exposed to PBDEs from 

mother’s milk (LaKind and Berlin 2000; Lindström 1999; Norön and Meironyté 2000; Schecter et al. 

2003b; Strandman et al. 2000; WHO 1994a). More data are needed on the levels of PBDEs exposure in 

nursing women, from occupational situations, from consumption of fish, and from those of the general 

population. Exposure and body-burden studies related to consumption of fish in the U.S. population are 

needed to determine exposure levels, particularly in children of recreational and subsistence fishers. 

Exposure and body-burden studies are also needed in Native American communities that consume high 

levels of game and marine mammals. Information related to the exposure of children living near 

hazardous waste sites is also needed. In particular, information is needed that is related to the potential 

for children to be exposed to PBDEs bound to soil and dust particles through pica or unintentional hand-
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8. POTENTIAL FOR HUMAN EXPOSURE 

to-mouth activity within homes located in these areas. Quantitative information regarding the 

bioavailability and amount of PBDEs that children are exposed to through contact with contaminated 

soils is unavailable. Therefore, any information concerning this subject would be useful in evaluating 

children’s exposure. Additional information on weight-adjusted intakes would be helpful for determining 

the health risks for young children, particularly those in Native American populations. Infants and young 

children consume a greater amount of food per kilogram of body weight and, therefore, may have a 

proportionately greater exposure to PBDEs than adults. 

Child health data needs relating to susceptibility are discussed in Section 5.12.2 Identification of Data 

Needs: Children’s Susceptibility. 

Exposure Registries. 

Polybrominated Biphenyls. The Michigan Department of Community Health (MDCH), together with the 

Centers for Disease Control and Prevention (CDC) and three other federal agencies, began a major study 

to assess the health effects of PBBs after the Michigan contamination episode. A health questionnaire 

and blood samples were collected from people affected by the feed-contamination incident. MDCH had 

the responsibility to analyze several thousand samples for PBB from 1975 to 1978. MDCH continues 

contact with this cohort, updates health questionnaires, and collects blood samples to be analyzed (MDCH 

2002). 

Polybrominated Diphenyl Ethers. No exposure registries for PBDEs were located. This substance is not 

currently one of the compounds for which a subregistry has been established in the National Exposure 

Registry. The substance will be considered in the future when chemical selection is made for 

subregistries to be established. The information that is amassed in the National Exposure Registry 

facilitates the epidemiological research needed to assess adverse health outcomes that may be related to 

exposure to this substance. 

8.8.2 Ongoing Studies 

Polybrominated Biphenyls. A search in Federal Research in Progress (FEDRIP 2002) did not identify 

ongoing research studies for PBBs. 
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8. POTENTIAL FOR HUMAN EXPOSURE 

Polybrominated Diphenyl Ethers. A search of FEDRIP (2003) did not identify ongoing research studies 

for PBDEs. However, EPA has funded work in the Science To Achieve Results (STAR) Research Grants 

program, as well as regional and intramural efforts (EPA 2004). In addition, the National Toxicology 

Program (NTP) has committed to performing studies on the commercial pentaBDE and octaBDE 

products and their congener components (NTP 2004). 
PBBs and PBDEs 467 

9. ANALYTICAL METHODS 

The purpose of this chapter is to describe the analytical methods that are available for detecting, 

measuring, and/or monitoring polybrominated biphenyls (PBBs) and polybromintaed diphenyl ethers 

(PBDEs), its metabolites, and other biomarkers of exposure and effect to PBBs and PBDEs. The intent is 

not to provide an exhaustive list of analytical methods. Rather, the intention is to identify wellestablished methods that are used as the standard methods of analysis. Many of the analytical methods 

used for environmental samples are the methods approved by federal agencies and organizations such as 

EPA and the National Institute for Occupational Safety and Health (NIOSH). Other methods presented in 

this chapter are those that are approved by groups such as the Association of Official Analytical Chemists 

(AOAC) and the American Public Health Association (APHA). Additionally, analytical methods are 

included that modify previously used methods to obtain lower detection limits and/or to improve accuracy 

and precision. 

PBBs and PBDEs are analyzed in environmental and biological samples by methods quite similar to those 

used for polychlorinated biphenyls (PCBs) (de Kok et al. 1977; Fries 1985b; Pomerantz et al. 1978). The 

analytical methods for PBBs were developed primarily in the 1970s, whereas analytical methods for 

PBDEs were developed very recently. In the 30 years between analytical efforts directed at PBBs and 

PBDEs, there have been many advances in the technology and costs of analytical instruments. Thus, 

while gas chromatography-electron capture detection (GC-ECD) with packed columns (i.e., noncongener 

specific) was the primary analytical technique used in the 1970s for PBBs, gas chromatography-mass 

spectrometry (GC-MS) with capillary columns (i.e., congener specific) is the primary analytical technique 

now used for PBDEs. These points should be considered when comparing the differences and quality of 

analysis for these two classes of compounds. 

Covaci et al. (2003) recently reviewed the determination of brominated flame retardants, with emphasis 

on PDBEs in environmental and human samples. The analysis methodology for PBBs and PBDEs 

includes several steps: sample collection and storage, sample pretreatment, extraction, cleanup and 

fractionation, and analytical determination. Care must be taken to assure that the sample collection 

follows quality-assurance protocols and that equipment and containers are free from contamination. It is 

important that laboratories utilize blanks when reporting trace levels of PBBs and PBDEs. This practice 

will minimize the influence of trace contamination samples that can originate from a variety of sources. 
PBBs and PBDEs 468 

9. ANALYTICAL METHODS 

Most sample collections are by grab sampling; however, PBBs and PBDEs may be concentrated from 

water onto sorbents. Desiccation of solid samples (e.g., soil, sediment, and sewage sludge) is largely 

done for convenience. Dry samples are more efficiently homogenized, allowing for parallel 

determination of other analytes (e.g., lipid content) (Covaci et al. 2003). 

PBBs and PBDEs are typically separated from the biological and environmental media by extraction with 

organic solvents. Liquid-solid extraction (e.g., Soxhlet apparatus) remains a widely used technique for 

solid samples despite recent advances in other extraction techniques. Typical solvents are hexane, 

toluene, hexane/acetone mixtures, or dichloromethane. New extractions techniques, such as accelerated 

solvent extraction (ASE) or microwave-assisted extraction (MAE), are also currently used by a number of 

laboratories. The advantage of these techniques is lower solvent consumption and reduced extraction 

time. Supercritical fluid extraction (SFE) with solid-phase trapping has been used for the extraction 

brominated flame retardants from sediment with CO2 as the supercritical fluid. Extraction with 

pressurized hot water (PHWE) has been used for the analysis of brominated analytes from sediment. 

Liquid-liquid extraction has been applied for river and seawater samples, using hexane/acetone mixtures. 

Solid-phase extraction (SPE) has been used for the analysis of acidic and neutral brominated flame 

retardants from human plasma (Covaci et al. 2003). 

Cleanup steps are necessary to remove compounds that may interfere with the determination (e.g., humic 

acids, lipids) of PBBs and PBDEs. Lipids (e.g., oils and fats) may be destroyed with concentrated 

sulfuric acid treatment either directly to the extract or using impregnated silica columns. 

Chromatography (e.g., gel permeation, silica gel, Florisil) is used to remove other matrix interferences 

and to fractionate samples (Covaci et al. 2003). 

The identification and quantitation of PBBs and PBDEs are most often accomplished by GC techniques. 

Capillary or high-resolution gas chromatography (HRGC) columns capable of separating a substantial 

proportion of the congeners are indispensable, and GC detectors possessing high selectivity and 

sensitivity for the PBBs and PBDEs are required. The more universal and less sensitive flame-ionization 

detector (FID) is used much less often than the electron-capture detector (ECD), which has exceptional 

sensitivity to highly brominated compounds. The mass-spectrometer detectors have sensitivities 

somewhat lower than ECD, and they have even greater selectivity for PBBs and PBDEs and can 

distinguish and individually measure homologs that may co-elute on a particular HRGC column. The use 

of MS is indispensable in the definitive identification of PBB and PBDE congeners. A recent method of 

detection is electron-capture negative ionization (ECNI) as an ionization technique in combination with 
PBBs and PBDEs 469 

9. ANALYTICAL METHODS 

GC-MS analysis (de Boer et al. 2000). This method is advantageous because it offers a high sensitivity 

for compounds with four or more bromine atoms. The sensitivity is approximately 10 times higher than 

with the use of the ECD. However, ECNI, although generally more sensitive and less costly than other 

ionization methods for PBDE analysis, does not provide information on the molecular ion cluster (as 

required for qualitative identification). It is also more subject to brominated interferences and does not 

allow the used of 13C-labeled standards for quantification (Ikonomou and Rayne 2002). Conversely, 

electron ionization (EI) methods suffer from fragmentation of the molecular ions, creating difficulties in 

both identification and quantitation of congeners in full-scan and single ion monitoring (SIM) modes, 

respectively. For example, loss of Br atoms from PBDE congeners during EI may lead to incorrect 

identification of the parent ion as a lower brominated congener. In addition, the relatively unpredictable 

fragmentation during EI or ECD restricts the utility of applying relative response factors (RRFs) of one 

congener for which an analytical standard is available (e.g., 2,2,4,4’-tetrabromodiphenyl ether or BDE 47) 

for other members of its homolog group (e.g., tetrabromodiphenyl ether [tetraBDEs]). This can result in 

either under- or overestimating concentrations of congeners for which analytical standards are not 

available (Ikonomou and Rayne 2002b). In general, hepta- through deca-BDE congeners are difficult to 

determine accurately by GC analysis, especially in biological samples (Ikonomou and Rayne 2002b). 

The analysis of decaBDE (i.e., BDE 209) and 2,2’,4,4’,5,6’-hexabromodiphenyl ether (i.e., BDE 154) has 

some analytical difficulties. For example, BDE 209 (1) is not stable at high temperatures in the GC 

injector and GC column; (2) is sensitive to degradation by UV light (i.e., both sunlight and fluorescent 

light); (3) behaves differently in the MS source from those of chlorinated and lower brominated 

compounds (de Boer and Cofino 2002); and (4) may easily adsorb to small dust particles in the 

laboratory, which may result in sample contamination (Covaci et al. 2003). Thermal decomposition of 

BDE 209 can be avoided using a short GC column and a thermally inert GC injection port. In contrast, 

BDE 154 usually coelutes from most gas chromatographic columns with 2,2’,4,4’,5,5’-hexabromobiphenyl (PBB-153). In order to ensure the separation of BDE 154 and PBB 153, analysts need to use a 

sufficiently long GC column. Thus, in order to accurately determine the levels of BDE 209 and BDE 154 

in analytical samples, analysts are required perform two separate GC measurements under different 

operating conditions. 

9.1 BIOLOGICAL MATERIALS 

Methods for the determination of organobromine compounds such as PBBs and PBDEs generally consist 

of the following steps: extraction of the analyte from the sample matrix; cleanup to remove interfering 
PBBs and PBDEs 470 

9. ANALYTICAL METHODS 

compounds; and analysis (separation and quantitation). The primary method of analysis is GC coupled 

with ECD or MS. Analytical methods have been developed for the determination of PBBs and PBDEs in 

blood or serum, urine, feces, adipose tissue, liver, and breast milk. The methods for determining PBB and 

PBDE residues in biological samples are given in Tables 9-1 and 9-2, respectively. 

Polybrominated Biphenyls. Residues in biological samples can be extracted using hexane/ether, 

petroleum ether/diethyl ether, toluene/ethyl acetate, or methylene chloride (Burse et al. 1980; Domino et 

al. 1980; Fawkes et al. 1982; Fehringer 1975b; Wolff et al. 1979b). Elution of samples on a florisil 

column, which is used for the cleanup of extracts with petroleum ether, separates PBBs from interfering 

substances (Pomerantz et al. 1978). As in the case of PCBs, the solvent(s) used for the extraction of a 

sample and the method used for the cleanup of an extract is dependent on the sample matrix (Pomerantz 

et al. 1978). Quantitation is usually done by GC. The major difference between the methods for the 

determination of PCBs and PBBs arises from the lower volatility of PBBs compared to PCBs. Due to the 

lower volatility of PBBs, the GC method is performed at a higher temperature and low liquid-phase load 

of the stationary phase. Capillary columns are required for the separation of the individual congeners in a 

mixture (Robertson et al. 1983b). However, decabromobiphenyl is so nonvolatile that a very short 

capillary column and high carrier gas linear velocity are required, which reduces the advantage of the 

capillary column over the packed column (Farrell 1980). Peaks from individual congeners of PBBs are 

detected and quantified with ECD (Robertson et al. 1983b). In general, retention time in gas 

chromatographic columns and response of ECD increase with increasing bromination. PBB residues in a 

sample can be confirmed by thin-layer chromatography, photochemical-alteration method, halogenspecific gas-chromatographic detection, or MS (de Kok et al. 1977; Erney 1975; Pomerantz et al. 1978). 

High recoveries (80–90%) of PBB residues are obtained by the available analytical methods. Typically, 

the limit of quantitation for PBB residues is about 1 μg/kg in blood serum, 1 μg/kg in human milk, and 

0.5 μg/kg in adipose tissue (Eyster et al. 1983; Wolff et al. 1979a). An interlaboratory study is available 

that validates the precision and accuracy of PBB residue determination in human serum by a commonly 

used method (Burse et al. 1980). 

Polybrominated Diphenyl Ethers. Residues in biological samples can be extracted using sulfuric acid, 

2-propanol/hexane, methylene chloride, n-hexane, formic acid/2-propanol/water, or hexane/methyl t-butyl 

ether (Cramer et al. 1990; Meironyté Guvenius et al. 2001; Ohta et al. 2002a; Sellström et al. 1993; 

Sjödin et al. 1999; Thomsen et al. 2001b). Samples are cleaned up to remove interferences using Florisil, 

silica gel, alumina or activated-charcoal column chromatography, gel-permeation chromatography (GPC), 
PBBs and PBDEs 471 

9. ANALYTICAL METHODS 

Table 9-1. Analytical Methods for Determining PBBs in Biological Materials 

Sample 

matrix Preparation method 

Analytical 

method 

Sample 

detection limit

Percent 

recovery Reference 

Serum Extract denatured sample with 

hexane-ethyl ether; clean up by 

Florisil column chromatography 

GC-ECD 1 μg/L 100.6–106.8 

at 100 μg/L 

Burse et al. 

1980 

Serum Extract denatured sample with 

hexane-ether; clean up by Florisil 

column chromatography 

GC-ECD 1 ng/g 86–92 Wolff et al. 1979

Plasma Extract denatured sample with 

petroleum ether-ethyl ether; clean 

up by Florisil and silica gel column 

chromatography 

GC-ECD 1.0 μg/L (for 

hexa) 

102 (for 

hexa) 

Willet et al. 1978

Whole 

blood 

Extract denatured sample with 

petroleum ether-ethyl ether; clean 

up by Florisil column chromatography 

GC-ECD 0.7 ng/g 90–96 Domino et al. 

1980 

Feces Extract sample with petroleum 

ether-ethyl ether; clean up by 

Florisil and silica gel column 

chromatography 

GC-ECD 1.4 ng/g (for 

hexa) 

61 (for hexa) Willet et al. 1978

Bile Extract denatured sample with 

petroleum ether-ethyl ether; clean 

up by Florisil and silica gel column 

chromatography 

GC-ECD 0.08 ng/g (for 

hexa) 

92 (for hexa) Willet et al. 1978

Milk Extract denatured sample with 

petroleum ether-ethyl ether; clean 

up by Florisil and silica column 

chromatography 

GC-ECD 1.4 μg/L (for 

hexa) 

108 (for 

hexa) 

Willet et al. 1978

Milk, 

human 

Extract with potassium oxalate, 

ethanol/diethyl ether, or hexane 

GC-ECD 1 ng/g No data Eyster et al. 

1983 

Liver Extract sample with methanolchloroform; clean up by acidic 

silica column chromatography 

GC-ECD No data 70 (for hexa) Fawkes et al. 

1982 

Adipose 

tissue 

Extract sample with methylene 

chloride; clean up by acidic silica 

gel column chromatography 

GC-ECD No data 80 Fawkes et al. 

1982 

Adipose 

tissue 

(exposed 

workers) 

Toluene/ethyl acetate (1+3); clean 

up using GPC/Bio beads 

GC-ECD 0.5 ng/g 98 Wolff et al. 1979

Human 

tissues 

(postmortem) 

Extract with hexane; clean up 

using Florisil column 

GC-ECD 0.5 ng/g No data Micelli et al. 

1985 

EC = electron capture detection; GC = gas chromatography; GPC = gel permeation chromatography; hexa = hexabrominated biphenyl; PBBs = polybrominated biphenyls; SIM = selected ion monitoring 
PBBs and PBDEs 472 

9. ANALYTICAL METHODS 

Table 9-2. Analytical Methods for Determining PBDEs in Biological Materials 

Sample 

matrix Preparation method 

Analytical 

method 

Sample 

detection limit 

Percent 

recovery Reference 

Animal 

tissues 

(muscle, fat, 

and egg) 

Extraction with sulfuric 

acid; clean up with GPC/

silca column/carbon 

column 

GC-MS 

(NCI) 

No data No data Sellström et 

al. 1993 

Human 

adipose 

tissue 

Soxhlet extraction; clean 

up using 2 solid-phase 

extraction cartridges 

Capillary 

GC-EILRMS 

0.05–0.30 ng/g 

lipid 

81–103 Covaci et al. 

2002 

Human 

adipose 

tissue 

Extraction with methylene 

chloride; evaporate; clean 

up on silica gel followed by 

clean up on alumina and 

on a carbon/silica gel 

column 

HRGC/

HRMS 

0.73–120 pg/g No data Cramer et al. 

1990 

Human 

liver/

adipose 

tissue 

Extract with 2-propanol/

hexane; clean up with 

Lipidex 5000, column 

chromatography/GPC 

GC-MS 

(NCI) 

5 pg/g lipids 83 (54–116) liver; 

71 (51–95) 

adipose 

Meironyte 

Guvenius et 

al. 2001 

Human milk Extract with potassium 

oxalate/ethanol/diethyl 

ether/pentane; GPC; clean 

up on Florisil; elute with 

hexane 

GC/MS 

(NCI/SIM) 

<0.6 ng/g fat No data WHO 1994a 

Human milk Extract by column 

chromatography using 

hexane/dichloromethane/

hexane; clean up using 

GPC 

GC-MS 

(SIM) 

5 pg/g lipids 86–102 Meironyté et 

al. 1999a, 

1999b 

Human milk Extract with n-hexane; 

clean up using multi-layer 

column 

HRGCLRMS or 

LRGCHRMS (EISIM) 

No data >80 Ohta et al. 

2002 

Human 

plasma 

Extract with formic acid, 

2-propanol, and water on a 

SPE column; derivatized 

using diazomethane 

GC-MS 

(NCI) 

1–10 pg/g 

plasma 

72 Thomsen et 

al. 2001b 

Human 

serum 

Extraction with hexane/

MTBE (1:1); clean up silica 

gel/surfuric acid column 

GC-ECD; 

GC-MS 

(NCI) 

0.7 ng/g lipid 

weight 

69–104 (low 

spike); 77–104 

(high spike) 

Sjödin et al. 

1999a 

ECD = electron capture detection; EI = electron impact; EILR = electron impact low-resolution; GC = gas 

chromatography; GPC = gel permeation chromatography; HRGC = high resolution gas chromatography; 

HRMS = high resolution mass spectrometry; LRGC = low resolution gas chromatography; LRMS = low resolution 

mass spectrometry; MS = mass spectrometry; MTBE = methyl-tert-butyl either; NCI = negative chemical ionization; 

PBDEs = polybrominated diphenyl ethers; SIM = selected ion monitoring; SPE = solid phase extraction 
PBBs and PBDEs 473 

9. ANALYTICAL METHODS 

and/or liquid chromatography (LC) (Sellström et al. 1993; Cramer et al. 1990; Meironyté Guvenius et al. 

2001; Sellström et al. 1993; Sjödin et al. 1999). Most techniques are based on analysis by GC with ECD 

or coupled with MS (WHO 1994a). Capillary columns and temperature programming allow the 

separation of the different PBDE congeners. High recoveries (69–104%) of PBDE residues are obtained 

by the available analytical methods. Typically, the limit of quantitation for PBDE residues is about 

0.7 ng/g lipid in blood serum, 5 pg/g lipid in human milk, and 0.3 ng/g lipid in adipose tissue (Covaci et 

al. 2002b; Meironyté Guvenius 1999a, 1999b; Sjödin et al. 1999). 

9.2 ENVIRONMENTAL SAMPLES 

Most environmental analyses have been performed using multiresidue methods involving solvent 

extraction of the analytes from the sample matrix, cleanup to remove interfering compounds, 

determination by GC with ECD, and confirmation using an ancillary method such as MS. New methods 

and technologies are evolving, and this has resulted in lower detection limits. For example, detection 

limits for PBBs are in the low parts-per-billion (ppb) to parts-per-trillion (ppt) range for water matrices 

and in the low parts-per-million (ppm) to ppb range for food; for PBDEs, detection limits are in the low 

ppb range for water matrices and in the low ppb to ppm range for fish tissues. Analytical methods for the 

determination of PBBs and PBDEs in environmental samples are given in Tables 9-3 and 9-4, 

respectively. 

Polybrominated Biphenyls. Residues in environmental samples can be extracted using hexane-ether, 

petroleum ether-ether, toluene-ethyl acetate, or methylene chloride (Burse et al. 1980; Domino et al. 

1980; Fawkes et al. 1982; Fehringer 1975b; Wolff et al. 1979b). As for biological samples, quantitation 

of environmental samples is also usually done by GC. Capillary columns are required for the separation 

of the individual congeners in a mixture (Robertson et al. 1983b). High recoveries (74–98%) of PBB 

residues in environmental samples are obtained by the available analytical methods. Typically, the limit 

of quantitation for PBB residues is about 0.1 μg/kg in soil and 0.7 μg/kg in sediment (Jacobs et al. 1976, 

1978; Kuosmanen et al. 2002). 

Polybrominated Diphenyl Ethers. Like PCBs, air samples containing PBDEs are usually collected by 

pumping air through a sampler containing a glass-fiber filter and adsorbent trap to separate the particlebound and vapor-phase fractions, respectively (Dobber et al. 2000a; Hillery et al. 1997). The filters and 

adsorbents are then Soxhlet extracted with acetone/hexane, and the extracts are cleaned up and analyzed 

by high-resolution GC techniques. 
PBBs and PBDEs 474 

9. ANALYTICAL METHODS 

Table 9-3. Analytical Methods for Determining PBBs in Environmental Samples 

Sample 

Matrix Preparation method 

Analytical 

method 

Sample 

detection limit

Percent 

recovery Reference 

Commercial Sample dissolved in benzene 

FireMaster BP-6 

GC-ECD 1.6 ng (EC) 

GC-PED 

Not applicable 2.8 ng 

(PED) 

Mulligan et al. 

1980 

Soil Extract sample with hexaneacetone; clean up by Florisil column 

chromatography 

GC-ECD 0.1 ng/g 74.2–83.2 

(for hexa) 

Jacobs et al. 

1976, 1978 

Soil Extraction using hexane/acetone; 

clean up using Florisil column 

GCFID/ECD No data No data Hill et al. 1982

Plant tissue Extract macerated sample with 

hexane-acetone; clean up by Florisil 

column chromatography 

GC-ECD 0.3 ng/g No data Jacobs et al. 

1978 

Effluent and 

river water 

Extract sample with hexane-ethyl 

ether 

GC-ECD 0.1 ng/g 90 Hesse and 

Powers 1978 

Sediment Extract sample with hexane-acetoneGC-ECD No data No data Hesse and 

Powers 1978 

Sediment Pressurized hot water extraction 

coupled with clean up by LC 

LC-GC-MS/

FID 

0.71 ng/g No data Kuosmanen et 

al. 2002 

Fish Extract homogenized sample with 

hexane-water; clean up by acidic 

and basic silica columns 

GC-ECD No data 98 (for 

hexa) 

Gobas et al. 

1989 

Fish Extract homogenized sample with 

hexane-methylene chloride; clean 

up by gel permeation and silica gel 

chromatography 

HRGCHRMS No data No data Kuehl et al. 

1991 

Fish Extract homogenized sample with 

hexane-acetone; clean up by gel 

permeation chromatography 

HRGC-MS/

NCI and 

HRGC-ECD

No data No data Jaffe et al. 

1985 

Terrestrial, 

fresh water, 

and marine 

samples 

Extraction with diethyl ether/hexane; 

hydrolysis with 98% sulfuric acid/bio 

beads/silica gel/activated charcoal 

MS (NCI) No data No data Jansson et al. 

1991, 1993 

Dolphin fat Soxhlet extraction using hexanemethylene chloride; clean up using 

GPC, silica gel 

MS No data No data Kuehl et al. 

1991 

Animal feeds Elute ground sample containing 

celite with methylene chloride; clean 

up by Florisil column 

chromatography 

GC-ECD 8 ng/g (for 

hexa) 

98 (for 

hexa) 

Fehringer 

1975b 

Dairy 

products 

Fat extracted by methanol/ether; 

clean up by GPC, 25% toluene in 

ethyl acetate 

GC-ECD 7 ng/g No data Fehringer 

1975 

Plants Cut, extracted with hexane/acetone; 

clean up with Florisil column 

GC-ECD 0.3 ng/g wet 

basis 

No data Chou et al. 

1978 

EC = electron capture detection; FID = flame ionization detector; GC = gas chromatography; hexa = hexabrominated 

biphenyl; HRGC = high resolution gas chromatography; HRMS = high resolution mass spectrometry; LC = liquid 

chromatography; MS = mass spectrometry; NCI = negative chemical ionization; PED = plasma emission detection; 

PBBs = polybrominated biphenyls 
PBBs and PBDEs 475 

9. ANALYTICAL METHODS 

Table 9-4. Analytical Methods for Determining PBDEs in Environmental Samples

Sample 

Matrix Preparation method 

Analytical 

method 

Sample 

detection limit 

Percent 

recovery Reference 

Air Air pumped through glass 

fiber filter and adsorbent trap; 

filters and adsorbents are 

Soxhlet extracted with 

acetone/hexane; cleaned-up 

by column chromatography 

GC/MS No data No data Dodder et al. 

2000a 

Water Clean up by disk-type C18 

solid-phase extraction 

Capillary 

GC-ECD 

0.12 pg/L 103±8.6 

(river water); 

87±10.7 (sea 

water) 

Yamamoto et 

al. 1997 

Sewage Extract with chloroform; 

evaporate and dissolve 

residue in ethanol 

GC/MS 0.06 μg/g No data WHO 1994a 

Sediment Clean up by cartridge-type 

Florosil extraction 

Capillary 

GC-ECD 

9.7 ng/g 91±6.3 Yamamoto et 

al. 1997 

Sediment Pressurized hot water 

extraction coupled with clean 

up by LC 

LC-GCMS/FID 0.71 ng/g No data Kuosmanen et 

al. 2002 

Sediment Extract with acetone; clean 

up on Florisil 

NAA; 

GC/EC 

<5 ng/g; 

<5 ng/g 

No data Watanabe et al. 

1987 

Fish Extract with acetone-hexane 

+ hexane-ethyl ether; treatment with sulfuric acid or 

clean up on alumina; 

chromatography on silica gel 

GC/EC; 

GC/MS 

0.1 μg/g fat No data Anderson and 

Blomkvist 1981 

Fish Extract with dichloromethane 

on chromatography column; 

clean-up using GPC; 

fractionation using silica gel 

column 

GC-HRMS 

(NCI) 

5–93 pg/g No data Alaee et al. 

2001 

Fish Extract clean up with GPC 

and mini-column chromatography; concentration 

GC-MS 

(NCI) 

0.01–0.2 ng/g 

lipid 

88–128 Akutsu et al. 

2001 

Animal 

tissues 

Homogenize; extract with 

n-hexane-acetone; treatment 

with sulfuric acid; GPC; 

chromatography or silica gel 

chromatography or activated 

charcoal 

GC/MS 

(NCI) 

10 pg/g No data Jansson et al. 

1991 

ECD = electron capture detection; GC = gas chromatography; GPC = gel permeation chromatography; 

HRMS = high resolution mass spectrometry; LC = liquid chromatography; MS = mass spectrometry; NAA = neutron 

activation analysis; NCI = negative chemical ionization; PBDEs = polybrominated diphenyl ethers 
PBBs and PBDEs 476 

9. ANALYTICAL METHODS 

Residues in environmental samples can be extracted using chloroform, acetone, acetone-hexane, hexaneacetone, and hexane-ether (Anderson and Blomkvist 1981; Jansson et al. 1991; Watanabe et al. 1987; 

WHO 1994a). Samples are cleaned up to remove interferences using Florisil, silica gel, alumina or 

activated charcoal column chromatography, gel permeation chromatography (GPC), and/or liquid 

chromatography (LC) (Akutsu et al. 2001; Alaee et al. 2001; Anderson and Blomkvist 1981; Jansson et 

al. 1991; Watanabe et al. 1987; Yamamoto et al. 1997). As for biological samples, quantitation of 

environmental samples is also usually done by GC. Capillary columns are required for the separation of 

the individual congeners in a mixture (WHO 1994a). High recoveries (88–128%) of PBDE residues in 

environmental samples are obtained by the available analytical methods (Akutsu et al. 2001). Typically, 

the limit of quantitation for PBDE residues is about 0.12 ng/mL in water, 9.7 μg/kg in sediment, and 

0.2 μg/kg lipid in fish (Akutsu et al. 2001; Yamamoto et al. 1997). The first inter-laboratory study on 

PBDEs in environmental samples showed that there is good agreement for quantification of BDE 47 and 

2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100) congeners. However, improved methods are required 

for analysis of 2,2’,4,4’,5-pentabromodiphenyl ether (BDE 99), 2,2’,4,4’,5,5’-hexabromodiphenyl ether 

(BDE 153), 2,2’,4,4’,5,6’-hexabromodiphenyl ether (BDE 154), and BDE 209 congeners (de Boer 2000). 

9.3 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the 

Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of PBBs and PBDEs is available. Where adequate information 

is not available, ATSDR, in conjunction with NTP, is required to assure the initiation of a program of 

research designed to determine the health effects (and techniques for developing methods to determine 

such health effects) of PBBs and PBDEs. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will be 

evaluated and prioritized, and a substance-specific research agenda will be proposed. 
PBBs and PBDEs 477 

9. ANALYTICAL METHODS 

9.3.1 Identification of Data Needs 

Methods for Determining Biomarkers of Exposure and Effect. 

Exposure. Methods used as biomarkers for exposure to PBBs and PBDEs are available (Brilliant et al. 

1978; Covaci et al. 2002b; Eyster et al. 1983; Landrigan et al. 1979; Meironyté Guvenius 1999a, 1999b; 

Sjödin et al. 1999; Wolff et al. 1982). Analytical methods of sufficient precision and accuracy are 

presently available for the determination of PBBs and PBDEs in adipose tissue, serum, and breast milk 

(Burse et al. 1980; Covaci et al. 2002b; Domino et al. 1980; Fawkes et al. 1982; Fehringer 1975a; 

Meironyté Guvenius 1999a, 1999b; Sjödin et al. 1999; Willet et al. 1978; Wolff et al. 1979a, 1979b). 

Additional congener standards are needed for PBB and PBDEs analysis. Only 30–40 congener standards 

are currently available for identification and quantification of PBDEs (Eljarrat et al. 2002; Sjödin et al. 

1998). Metabolites are also important biomarkers for exposure to PBBs and PBDEs. However, these 

compounds are mostly unknown, and standards are not available. 

Effect. No studies have been conducted to determine if known effects of PBBs and PBDEs exposure can 

be quantitatively correlated with PBB or PBDE exposure. 

Methods for Determining Parent Compounds and Degradation Products in Environmental 

Media. Analytical methods of sufficient sensitivity are presently available for the determination of 

PBBs and PBDEs in environmental samples (Akutsu et al. 2001; Anderson and Blomkvist 1981; Covaci 

et al. 2003; Fehringer 1975b; Hesse and Powers 1978; Jacobs et al. 1976, 1978; Yamamoto et al. 1997). 

It would be helpful to develop data determining the detection limit and accuracy of PBBs determinations 

in fish and other aquatic animals (e.g., seals) and in sediment (Gobas et al. 1989; Jaffe et al. 1985; Kuehl 

et al. 1991). Analytical methods for determining lower brominated PBBs in environmental samples are 

available (Morris et al. 1992). An analytical method to determine PBB metabolites in fish would be 

helpful. A method for determining of 6-hydroxy-2,2',4,4',5,5'-hexabromobiphenyl, a metabolite of 

2,2',4,4',5,5'-hexabromobiphenyl, in dog feces is available (Gardner et al. 1979). Photochemical 

degradation leads to the formation of lower brominated products, which are the only environmental 

degradation products identified for PBBs. Analytical methods are presently available for the 

determination of these compounds in environmental samples (De Kok et al. 1977; Hill et al. 1982; 

Robertson et al. 1983b). There is no evidence in the literature of detectable biodegradation of PBBs in the 

environment under aerobic conditions (Griffin and Chou 1981a, 1981b), but the compounds may 
PBBs and PBDEs 478 

9. ANALYTICAL METHODS 

biodegrade to debrominated products under anaerobic conditions in polluted environments (Morris et al. 

1992). 

It would be helpful to develop data determining the accuracy of PBDE determinations (e.g., percent 

recovery) in environmental samples. Methods for determining degradation products and metabolites of 

PBDE are needed. There is no information in the literature of detectable biodegradation of PBDEs in the 

environment under aerobic or anaerobic conditions. The analysis of PBDE pyrolysis degradation 

products, such as polybrominated dibenzo-p-dioxins and dibenzofurans (PBDD/Fs), is often disturbed by 

the presence of PBDEs. Ebert et al. (1999) demonstrated that by using a Florisil column in a sample 

clean-up process, almost complete separation of PBDEs and PBDDs/PBDFs is achieved before analysis 

by GC-MS. 

9.3.2 Ongoing Studies 

No ongoing studies regarding analytical methods for determining PBBs and PBDEs residues or 

metabolites were located as a result of a search of the Federal Research in Progress Database (FEDRIP 

2003). 
PBBs and PBDEs 479 

10. REGULATIONS AND ADVISORIES 

Table 10-1 summarizes international, national, and state regulations and guidelines on human exposure to 

PBBs and PBDEs. 

ATSDR has derived an MRL of 0.01 mg/kg/day for acute-duration oral exposure to PBBs. The MRL is 

based on a NOAEL for thyroid effects in rats (Allen-Rowland et al. 1981). Intermediate- and chronicduration oral MRLs were not derived because serious developmental effects (fetal abortion and stillbirths) 

were observed in monkeys that had been exposed to PBBs for durations that spanned the intermediate and 

chronic categories at the lowest dose tested in the database. This dose, 0.012 mg/kg/day, caused 

increased menstrual cycle duration and implantation bleeding after 6–7 months of exposure and fetal 

deaths after1 year of exposure in monkeys, with surviving infants having decreased birth weight and 

decreased postnatal weight gain and weight loss also occurring in maternal animals (Allen et al. 1978, 

1979; Lambrecht et al. 1978). The reproductive effects are less serious, but concern for serious 

developmental toxicity following exposures of >1 year precludes deriving an MRL for intermediateduration exposure. Derivation of an MRL for chronic oral exposure is precluded by the serious 

developmental effects that occurred following exposures exceeding 1 year in duration. 

ATSDR has derived an MRL of 0.03 mg/kg/day for acute-duration oral exposure to lower brominated 

diphenyl ethers (BDEs). The acute oral MRL is based on a NOAEL of 1 mg/kg/day for reduced serum 

levels of thyroid T4 hormone in fetal rats that were exposed to a commercial pentaBDE mixture on 

days 4–20 of gestation (Zhou et al. 2002). ATSDR also derived an MRL of 0.007 mg/kg/day for 

intermediate-duration oral exposure to lower brominated BDEs. The intermediate oral MRL is based on a 

minimal LOAEL of 2 mg/kg/day for liver effects in rats that were exposed to a commercial pentaBDE 

mixture for 90 days (WIL Research Laboratories 1984). A chronic-duration oral MRL was not derived 

for lower brominated BDEs due to insufficient data. 

ATSDR has derived an MRL of 10 mg/kg/day for intermediate-duration oral exposure to decaBDE. This 

intermediate oral MRL is based on a NOAEL of 1000 mg/kg/day for developmental toxicity in rats 

exposed to decaBDE for 19 days during gestation (Hardy et al. 2002). No acute- or chronic-duration oral 

MRLs were derived for decaBDE due to insufficient data. 

ATSDR has derived an MRL of 0.006 mg/m3

 for intermediate-duration inhalation exposure to lower 

brominated BDEs. The intermediate inhalation MRL is based on a NOAEL of 1.1 mg/m3

 for thyroid 
PBBs and PBDEs 480 

10. REGULATIONS AND ADVISORIES 

Table 10-1. Regulations and Guidelines Applicable to PBBs and PBDEsa

Agency Description Information Reference 

INTERNATIONAL 

Guidelines: 

 IARC Carcinogenicity classification IARC 1987 

 Decabromobiphenyl Group 2Bb

 

 Decabromodiphenyl ether Group 3c

 

 Hexabromobiphenyl Group 2Bb

 

 Octabromobiphenyl Group 2Bb

 

NATIONAL 

Regulations and Guidelines: 

a. Air 

 ACGIH TLV (8-hour TWA) No data 

 NIOSH REL (10-hour TWA) No data 

 OSHA PEL (8-hour TWA) No data 

b. Water 

 EPA Standards for owners and operators of 

hazardous waste TSD facilities; 

groundwater monitoring for p-bromodiphenyl ether 

 EPA 2002e 

 Suggested EPA method 8270 40CFR264, 

 PQL 10 μg/L Appendix IX 

d. Food 

 FDA Indirect food additives; substances for 

use only as components of adhesives 

 FDA 2001 

21CFR175.105 

(c)(5) 

 Hexabromo-1,1'-biphenyl 

e. Other 

 EPA Carcinogenicity classification IRIS 2002 

 Decabromodiphenyl ether Group Cd

 

 Nonabromodiphenyl ether Group De

 

 Octabromodiphenyl ether Group De

 

 Hexabromodiphenyl ether Group De

 

 Pentabromodiphenyl ether Group De

 

 Tetrabromodiphenyl ether Group De

 

 Tribromodiphenyl ether Group De

 

 p,p’-Dibromodiphenyl ether Group De

 

 p-Bromodiphenyl ether Group De

 

 EPA Oral reference dose (RfD) IRIS 2002 

 Decabromodiphenyl ether 1x10-2 mg/kg/day 
PBBs and PBDEs 481 

10. REGULATIONS AND ADVISORIES 

Table 10-1. Regulations and Guidelines Applicable to PBBs and PBDEsa

Agency Description Information Reference 

 Octabromodiphenyl ether 3x10-3 mg/kg/day 

NATIONAL (cont.)

 EPA Pentabromodiphenyl ether 2x10-3 mg/kg/day 

 Chemical substances subject to 

proposed or final TSCA rules or orders 

Regulated under 

TSCA section(s)

EPA 1998 

 Decabromobiphenyl 5(a)(2) 

 Decabromodiphenyl ether 4 

 Hexabromo-1,1'-biphenyl 5(a)(2) 

 Octabromobiphenyl 5(a)(2) 

 Octabromodiphenyl ether 4 

 p-Bromodiphenyl ether 5(a)(2) 

 Pentabromodiphenyl ether 4 

 EPCRA Section 313; toxic chemical EPA 2001 

 Decabromodiphenyl ether 

 Identification and listing of hazardous 

waste; commercial chemical products, 

manufacturing chemical intermediates, 

or off-specification commercial 

chemical products are identified as 

toxic wastes; EPA hazardous waste 

number 

 EPA 2002b 

40CFR261.33(f) 

 p-Bromodiphenyl ether U030 

 Land disposal restrictions; universal 

treatment standards 

 EPA 2002c 

40CFR268.48(a) 

 p-Bromodiphenyl ether 

 Wastewater 

standard 

0.055 mg/L 

 Nonwastewater 

standard 

15 mg/kg 

 Municipal solid waste landfills; 

hazardous constituent 

Suggested 

methods

PQL EPA 2002a 

40CFR258, 

Appendix II 

 p-Bromodiphenyl ether 8110 25 μg/L 

 8270 10 μg/L 

 Radiation protection; health and 

environmental protection standards for 

uranium and thorium mill tailings; listed 

constituent 

 EPA 2002d 

40CFR192, 

Appendix I 

 p-Bromodiphenyl ether 
PBBs and PBDEs 482 

10. REGULATIONS AND ADVISORIES 

Table 10-1. Regulations and Guidelines Applicable to PBBs and PBDEsa

Agency Description Information Reference 

NATIONAL (cont.)

 EPA CERCLA hazardous substance under 

Section 307(a) of the Clean Water Act 

and RCRA Section 3001; reportable 

quantity 

 EPA 2002f 

40CFR302.4 

 p-Bromodiphenyl ether 100 pounds 

 Chemicals subject to the Prior Informed 

Consent Procedure: International rightto-know 

 EPA 2002g 

 Decabromobiphenyl 

 Hexabromobiphenyl 

 Octabromobiphenyl 

 Toxic chemical release reporting; 

community right-to-know; effective date 

for reporting 

 EPA 2002h 

40CFR372.65(c) 

 Decabromodiphenyl ether 01/01/87 

 Polybrominated biphenyls 01/01/87 

 TSCA; chemical information rules; 

manufacturers and importers must 

submit a Preliminary Assessment 

Information Manufacturers’ Report for 

each site at which they manufacture or 

import each substance by the reporting 

date shown 

 EPA 2002i 

40CFR712.30(d) 

 Decabromodiphenyl ether 03/12/90 

 Octabromodiphenyl ether 03/12/90 

 Pentabromodiphenyl ether 03/12/90 

 TSCA; chemical substances required to 

be tested 

Chemical substances 

known to be manufactured between 01/01/84 

and date of 

promulgation 

EPA 2002j 

40CFR766.25 

(a)(1) 

 Decabromodiphenyl ether 

 Octabromodiphenyl ether 

 Pentabromodiphenyl ether 

 TSCA; health and safety data reportingf Effective 

date

Sunset 

date

EPA 2002l 

40CFR716.120 

(a) 

 Decabromodiphenyl ether 01/11/90 06/30/98 

 Octabromodiphenyl ether 06/30/98 01/11/90 

 Pentabromodiphenyl ether 01/11/90 06/30/98 
PBBs and PBDEs 483 

10. REGULATIONS AND ADVISORIES 

Table 10-1. Regulations and Guidelines Applicable to PBBs and PBDEsa

Agency Description Information Reference 

NATIONAL (cont.)

 EPA TSCA; reporting requirements including 

deadline for submitting the protocols 

 EPA 2002k 

40CFR766.35 

 Decabromodiphenyl ether 01/31/91 

 Octabromodiphenyl ether 01/31/91 

 Pentabromodiphenyl ether 02/06/95 

 EPA TSCA; significant new uses subject to 

reporting of p-bromodiphenyl ether 

Significant new use 

includes manufacturing, 

importing, or processing 

of 10,000 pounds or 

more per year per 

facility for any use 

EPA 2002m 

40CFR721.3430 

 TSCA; significant new uses subject to 

reporting of polybrominated biphenyl 

The significant new use 

is any use 

EPA 2002n 

40CFR721.1790 

 Decabromodiphenyl 

 Hexabromo-1,1'-biphenyl 

 Octabromobiphenyl 

 Voluntary children’s chemical 

evaluation program 

 EPA 2002o 

 Decabromodiphenyl ether 

 Octabromodiphenyl ether 

 Pentabromodiphenyl ether 

 NTP Carcinogenicity classification NTP 2002 

 Decabromobiphenyl Reasonably anticipated 

to be human 

carcinogens 

 Octabromobiphenyl 

 Polybrominated biphenyl 

STATE 

Regulations and Guidelines: 

a. Air No data 

b. Water 

 Florida Drinking water guideline 10 μg/L HSDB 2002 

 p-Bromodiphenyl ether 

c. Food No data 

d. Other 

 Florida Toxic substance BLR 2002 

 Polybrominated biphenyl 
PBBs and PBDEs 484 

10. REGULATIONS AND ADVISORIES 

Table 10-1. Regulations and Guidelines Applicable to PBBs and PBDEsa

Agency Description Information Reference 

STATE (cont.)

 Massachusetts Hazardous substances requires the 

labeling of containers of toxic 

substances in the workplace 

 BLR 2002 

 Decabromobiphenyl Extraordinary 

hazardousg

 Decabromdiphenyl ether 

 Hexabromobiphenyl Extraordinary 

hazardousg

 Hexabromo-1,1'-biphenyl Extraordinary 

hazardousg

 Octabromobdiphenyl Extraordinary 

hazardousg

 p-Bromodiphenyl ether 

 Polybrominated biphenyl 

 Michigan Critical materials register; requires all 

businesses discharging waste products 

into the water or any sewer system to 

report the annual amount 

 BLR 2002 

 p-Bromodiphenyl ether 

 Minnesota Hazardous substance BLR 2002 

 Decabromodiphenyl ether 

 Hexabromo-1,1'-biphenyl 

 New Jersey Hazardous substance; requires SIC 

employers to submit Community Rightto-Know survey listing environmental 

hazardous substances present at their 

facilities in quantities that exceed 

500 pounds 

 BRL 2002 

 Decabromodiphenyl ether 

 New York Hazardous substance BLR 2002 

 p-Bromodiphenyl ether 

 Reportable quantity (air) 100 pounds 

 Reportable quantity (land) 100 pounds 
PBBs and PBDEs 485 

10. REGULATIONS AND ADVISORIES 

Table 10-1. Regulations and Guidelines Applicable to PBBs and PBDEsa

Agency Description Information Reference 

STATE (cont.)

 Pennsylvania Hazardous substance; requires 

employers to complete the Hazardous 

Substance Survey Form annually 

 BLR 2002 

 Decabromodiphenyl ether Environmental hazard 

 p-Bromodiphenyl ether Environmental hazard 

 Polybrominated biphenyl Special hazard 

a

Polybrominated biphenyls category includes: decabromobiphenyl (CAS# 13654-09-6); decabromodiphenyl ether 

(CAS# 1163-19-5); hexabromobiphenyl (CAS# 59080-40-9); hexabromo-1,1'-biphenyl (CAS# 36355-01-8); 

hexabromodiphenyl ether (CAS# 36483-60-0); nonabromodiphenyl ether (CAS# 63936-56-1); octabromobiphenyl 

(CAS# 27858-07-7); octabromobiphenyl (CAS# 61288-13-9); octabromodiphenyl ether (CAS# 32536-52-0); 

p-bromodiphenyl ether (CAS# 101-55-3); p,p’-dibromodiphenyl ether (CAS# 2050-47-7); pentabromodiphenyl ether 

(CAS# 32534-81-9); polybrominated biphenyl (CAS# 59536-65-1); polybrominated biphenyl mixture (CAS# 67774-

32-7); tetrabromodiphenyl ether (CAS# 40088-47-9); and tribromodiphenyl ether (CAS# 49690-94-0). b

Group 2B: possibly carcinogenic to humans 

c

Group 3: not classifiable as to its carcinogenicity to humans 

d

Group C: possible human carcinogen 

e

Group D: not classifiable as to human carcinogenicity 

f

Health and safety data reporting: the listed chemical substances are subject to all provisions of 40CFR716. 

Manufacturers, importers, and processors of a listed substance are subject to the reporting requirements of 

Subpart A for that substance. 

g

Extraordinary hazardous and designated carcinogen 

ACGIH = American Conference of Governmental Industrial Hygienists; BRL = Business & Legal Reports, Inc.; 

CERCLA = Comprehensive Environmental Response Compensation and Liability Act; CFR = Code of Federal 

Regulations; EPA = Environmental Protection Agency; EPCRA = Emergency Planning and Community Right-toKnow Act; FDA = Food and Drug Administration; HSDB = Hazardous Substances Data Bank; IARC = International 

Agency for Research on Cancer; IRIS = Integrated Risk Information System; NIOSH = National Institute for 

Occupational Safety and Health; NTP = National Toxicology Program; OSHA = Occupational Safety and Health 

Administration; PEL = permissible exposure limit; PQL = practical quantitation limits; RCRA = Resource 

Conservation Recovery Act; REL = recommended exposure limit; RfD = oral reference dose; SIC = Standard 

Industrial Classification; TLV = threshold limit value; TSCA = Toxic Substances Control Act; TSD = treatment, 

storage, and disposal; TWA = time-weighted average 
PBBs and PBDEs 486 

10. REGULATIONS AND ADVISORIES 

effects in rats that were exposed to a commercial octaBDE mixture for 13 weeks (Great Lakes Chemical 

Corporation 2001a, 2001b). 

EPA derived reference doses (RfDs) for octaBDE and pentaBDE of 3x10-3 and 2x10-3 mg/kg/day, 

respectively (IRIS 2002). The current RfD for decaBDE is 1x10-2 mg/kg/day, but this is based on a study 

using a preparation that contained only 77% decaBDE; the RfD for decaBDE is currently under review by 

the IRIS program. 

The National Academy of Sciences (NRC 2000) derived a RfD of 4 mg/kg/day for decabromodiphenyl 

oxide based on a NOAEL of 1,120 mg/kg/day for liver pathology in rats that were exposed for 103 weeks 

(NTP 1986). 

IARC (1987) has classified decabromobiphenyl, hexabromobiphenyl, and octabromobiphenyl in 

Group 2B, possibly carcinogenic to humans. NTP (2002) has classified decabromobiphenyl, octabromobiphenyl, and polybrominated biphenyl as reasonably anticipated to be human carcinogens. EPA (IRIS 

2002) has classified decabromodiphenyl ether in Group D, as a possible human carcinogen, and nonabromodiphenyl ether, octabromodiphenyl ether, hexabromodiphenyl ether, pentabromodiphenyl ether, 

tetrabromodiphenyl ether, tribromodiphenyl ether, p,p’-dibromodiphenyl ether, and p-bromodiphenyl 

ether in Group D, not classifiable as to human carcinogenicity. 

The American Industrial Hygiene Association (AIHA) has established a Workplace Environmental 

Exposure Level (WEEL) of 5 mg/m3

 for decaBPE (AIHA 1996). 

Industry is required to report health and safety data for PBDEs, including decabromodiphenyl ether, 

octabromodiphenyl ether, and pentabromodiphenyl ether, under the Toxic Substances Control Act 

(TSCA) (EPA 2002l). 

The European Union has banned the sale of products containing more than 0.1% penta- and octaBDE 

effective August 15, 2004 (EU 2003b). In 2003, the state of California passed an identical ban, to go into 

effect January 1, 2008 (California Assembly 2003). A subsequent proposed bill would change the 

implementation date to June 1, 2006 (California Assembly 2004). 
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PBBs and PBDEs 559 

12. GLOSSARY 

Absorption—The taking up of liquids by solids, or of gases by solids or liquids. 

Acute Exposure—Exposure to a chemical for a duration of 14 days or less, as specified in the 

Toxicological Profiles. 

Adsorption—The adhesion in an extremely thin layer of molecules (as of gases, solutes, or liquids) to the 

surfaces of solid bodies or liquids with which they are in contact. 

Adsorption Coefficient (Koc)—The ratio of the amount of a chemical adsorbed per unit weight of 

organic carbon in the soil or sediment to the concentration of the chemical in solution at equilibrium. 

Adsorption Ratio (Kd)—The amount of a chemical adsorbed by sediment or soil (i.e., the solid phase) 

divided by the amount of chemical in the solution phase, which is in equilibrium with the solid phase, at a 

fixed solid/solution ratio. It is generally expressed in micrograms of chemical sorbed per gram of soil or 

sediment. 

Benchmark Dose (BMD)—Usually defined as the lower confidence limit on the dose that produces a 

specified magnitude of changes in a specified adverse response. For example, a BMD10 would be the 

dose at the 95% lower confidence limit on a 10% response, and the benchmark response (BMR) would be 

10%. The BMD is determined by modeling the dose response curve in the region of the dose response 

relationship where biologically observable data are feasible. 

Benchmark Dose Model—A statistical dose-response model applied to either experimental toxicological 

or epidemiological data to calculate a BMD. 

Bioconcentration Factor (BCF)—The quotient of the concentration of a chemical in aquatic organisms 

at a specific time or during a discrete time period of exposure divided by the concentration in the 

surrounding water at the same time or during the same period. 

Biomarkers—Broadly defined as indicators signaling events in biologic systems or samples. They have 

been classified as markers of exposure, markers of effect, and markers of susceptibility. 

Cancer Effect Level (CEL)—The lowest dose of chemical in a study, or group of studies, that produces 

significant increases in the incidence of cancer (or tumors) between the exposed population and its 

appropriate control. 

Carcinogen—A chemical capable of inducing cancer. 

Case-Control Study—A type of epidemiological study that examines the relationship between a 

particular outcome (disease or condition) and a variety of potential causative agents (such as toxic 

chemicals). In a case-controlled study, a group of people with a specified and well-defined outcome is 

identified and compared to a similar group of people without outcome. 

Case Report—Describes a single individual with a particular disease or exposure. These may suggest 

some potential topics for scientific research, but are not actual research studies. 

Case Series—Describes the experience of a small number of individuals with the same disease or 

exposure. These may suggest potential topics for scientific research, but are not actual research studies. 
PBBs and PBDEs 560 

12. GLOSSARY 

Ceiling Value—A concentration of a substance that should not be exceeded, even instantaneously. 

Chronic Exposure—Exposure to a chemical for 365 days or more, as specified in the Toxicological 

Profiles. 

Cohort Study—A type of epidemiological study of a specific group or groups of people who have had a 

common insult (e.g., exposure to an agent suspected of causing disease or a common disease) and are 

followed forward from exposure to outcome. At least one exposed group is compared to one unexposed 

group. 

Cross-sectional Study—A type of epidemiological study of a group or groups of people that examines 

the relationship between exposure and outcome to a chemical or to chemicals at one point in time. 

Data Needs—Substance-specific informational needs that if met would reduce the uncertainties of human 

health assessment. 

Developmental Toxicity—The occurrence of adverse effects on the developing organism that may result 

from exposure to a chemical prior to conception (either parent), during prenatal development, or 

postnatally to the time of sexual maturation. Adverse developmental effects may be detected at any point 

in the life span of the organism. 

Dose-Response Relationship—The quantitative relationship between the amount of exposure to a 

toxicant and the incidence of the adverse effects. 

Embryotoxicity and Fetotoxicity—Any toxic effect on the conceptus as a result of prenatal exposure to 

a chemical; the distinguishing feature between the two terms is the stage of development during which the 

insult occurs. The terms, as used here, include malformations and variations, altered growth, and in utero

death. 

Environmental Protection Agency (EPA) Health Advisory—An estimate of acceptable drinking water 

levels for a chemical substance based on health effects information. A health advisory is not a legally 

enforceable federal standard, but serves as technical guidance to assist federal, state, and local officials. 

Epidemiology—Refers to the investigation of factors that determine the frequency and distribution of 

disease or other health-related conditions within a defined human population during a specified period. 

Genotoxicity—A specific adverse effect on the genome of living cells that, upon the duplication of 

affected cells, can be expressed as a mutagenic, clastogenic, or carcinogenic event because of specific 

alteration of the molecular structure of the genome. 

Half-life—A measure of rate for the time required to eliminate one half of a quantity of a chemical from 

the body or environmental media. 

Immediately Dangerous to Life or Health (IDLH)—The maximum environmental concentration of a 

contaminant from which one could escape within 30 minutes without any escape-impairing symptoms or 

irreversible health effects. 

Incidence—The ratio of individuals in a population who develop a specified condition to the total 

number of individuals in that population who could have developed that condition in a specified time 

period. 
PBBs and PBDEs 561 

12. GLOSSARY 

Intermediate Exposure—Exposure to a chemical for a duration of 15–364 days, as specified in the 

Toxicological Profiles. 

Immunologic Toxicity—The occurrence of adverse effects on the immune system that may result from 

exposure to environmental agents such as chemicals. 

Immunological Effects—Functional changes in the immune response. 

In Vitro—Isolated from the living organism and artificially maintained, as in a test tube. 

In Vivo—Occurring within the living organism. 

Lethal Concentration(LO) (LCLO)—The lowest concentration of a chemical in air that has been 

reported to have caused death in humans or animals. 

Lethal Concentration(50) (LC50)—A calculated concentration of a chemical in air to which exposure for 

a specific length of time is expected to cause death in 50% of a defined experimental animal population. 

Lethal Dose(LO) (LDLO)—The lowest dose of a chemical introduced by a route other than inhalation 

that has been reported to have caused death in humans or animals. 

Lethal Dose(50) (LD50)—The dose of a chemical that has been calculated to cause death in 50% of a 

defined experimental animal population. 

Lethal Time(50) (LT50)—A calculated period of time within which a specific concentration of a 

chemical is expected to cause death in 50% of a defined experimental animal population. 

Lowest-Observed-Adverse-Effect Level (LOAEL)—The lowest exposure level of chemical in a study, 

or group of studies, that produces statistically or biologically significant increases in frequency or severity 

of adverse effects between the exposed population and its appropriate control. 

Lymphoreticular Effects—Represent morphological effects involving lymphatic tissues such as the 

lymph nodes, spleen, and thymus. 

Malformations—Permanent structural changes that may adversely affect survival, development, or 

function. 

Minimal Risk Level (MRL)—An estimate of daily human exposure to a hazardous substance that is 

likely to be without an appreciable risk of adverse noncancer health effects over a specified route and 

duration of exposure. 

Modifying Factor (MF)—A value (greater than zero) that is applied to the derivation of a Minimal Risk 

Level (MRL) to reflect additional concerns about the database that are not covered by the uncertainty 

factors. The default value for a MF is 1. 

Morbidity—State of being diseased; morbidity rate is the incidence or prevalence of disease in a specific 

population. 

Mortality—Death; mortality rate is a measure of the number of deaths in a population during a specified 

interval of time. 
PBBs and PBDEs 562 

12. GLOSSARY 

Mutagen—A substance that causes mutations. A mutation is a change in the DNA sequence of a cell’s 

DNA. Mutations can lead to birth defects, miscarriages, or cancer. 

Necropsy—The gross examination of the organs and tissues of a dead body to determine the cause of 

death or pathological conditions. 

Neurotoxicity—The occurrence of adverse effects on the nervous system following exposure to a 

chemical. 

No-Observed-Adverse-Effect Level (NOAEL)—The dose of a chemical at which there were no 

statistically or biologically significant increases in frequency or severity of adverse effects seen between 

the exposed population and its appropriate control. Effects may be produced at this dose, but they are not 

considered to be adverse. 

Octanol-Water Partition Coefficient (Kow)—The equilibrium ratio of the concentrations of a chemical 

in n-octanol and water, in dilute solution. 

Odds Ratio (OR)—A means of measuring the association between an exposure (such as toxic substances 

and a disease or condition) which represents the best estimate of relative risk (risk as a ratio of the 

incidence among subjects exposed to a particular risk factor divided by the incidence among subjects who 

were not exposed to the risk factor). An odds ratio of greater than 1 is considered to indicate greater risk 

of disease in the exposed group compared to the unexposed group. 

Organophosphate or Organophosphorus Compound—A phosphorus-containing organic compound 

and especially a pesticide that acts by inhibiting cholinesterase. 

Permissible Exposure Limit (PEL)—An Occupational Safety and Health Administration (OSHA) 

allowable exposure level in workplace air averaged over an 8-hour shift of a 40-hour workweek. 

Pesticide—General classification of chemicals specifically developed and produced for use in the control 

of agricultural and public health pests. 

Pharmacokinetics—The science of quantitatively predicting the fate (disposition) of an exogenous 

substance in an organism. Utilizing computational techniques, it provides the means of studying the 

absorption, distribution, metabolism, and excretion of chemicals by the body. 

Pharmacokinetic Model—A set of equations that can be used to describe the time course of a parent 

chemical or metabolite in an animal system. There are two types of pharmacokinetic models: data-based 

and physiologically-based. A data-based model divides the animal system into a series of compartments, 

which, in general, do not represent real, identifiable anatomic regions of the body, whereas the 

physiologically-based model compartments represent real anatomic regions of the body. 

Physiologically Based Pharmacodynamic (PBPD) Model—A type of physiologically based doseresponse model that quantitatively describes the relationship between target tissue dose and toxic end 

points. These models advance the importance of physiologically based models in that they clearly 

describe the biological effect (response) produced by the system following exposure to an exogenous 

substance. 
PBBs and PBDEs 563 

12. GLOSSARY 

Physiologically Based Pharmacokinetic (PBPK) Model—Comprised of a series of compartments 

representing organs or tissue groups with realistic weights and blood flows. These models require a 

variety of physiological information: tissue volumes, blood flow rates to tissues, cardiac output, alveolar 

ventilation rates, and possibly membrane permeabilities. The models also utilize biochemical information 

such as air/blood partition coefficients, and metabolic parameters. PBPK models are also called 

biologically based tissue dosimetry models. 

Prevalence—The number of cases of a disease or condition in a population at one point in time. 

Prospective Study—A type of cohort study in which the pertinent observations are made on events 

occurring after the start of the study. A group is followed over time. 

q1*—The upper-bound estimate of the low-dose slope of the dose-response curve as determined by the 

multistage procedure. The q1* can be used to calculate an estimate of carcinogenic potency, the 

incremental excess cancer risk per unit of exposure (usually μg/L for water, mg/kg/day for food, and 

μg/m3

 for air). 

Recommended Exposure Limit (REL)—A National Institute for Occupational Safety and Health 

(NIOSH) time-weighted average (TWA) concentrations for up to a 10-hour workday during a 40-hour 

workweek. 

Reference Concentration (RfC)—An estimate (with uncertainty spanning perhaps an order of 

magnitude) of a continuous inhalation exposure to the human population (including sensitive subgroups) 

that is likely to be without an appreciable risk of deleterious noncancer health effects during a lifetime. 

The inhalation reference concentration is for continuous inhalation exposures and is appropriately 

expressed in units of mg/m3 or ppm. 

Reference Dose (RfD)—An estimate (with uncertainty spanning perhaps an order of magnitude) of the 

daily exposure of the human population to a potential hazard that is likely to be without risk of deleterious 

effects during a lifetime. The RfD is operationally derived from the no-observed-adverse-effect level 

(NOAEL-from animal and human studies) by a consistent application of uncertainty factors that reflect 

various types of data used to estimate RfDs and an additional modifying factor, which is based on a 

professional judgment of the entire database on the chemical. The RfDs are not applicable to 

nonthreshold effects such as cancer. 

Reportable Quantity (RQ)—The quantity of a hazardous substance that is considered reportable under 

the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA). Reportable 

quantities are (1) 1 pound or greater or (2) for selected substances, an amount established by regulation 

either under CERCLA or under Section 311 of the Clean Water Act. Quantities are measured over a 

24-hour period. 

Reproductive Toxicity—The occurrence of adverse effects on the reproductive system that may result 

from exposure to a chemical. The toxicity may be directed to the reproductive organs and/or the related 

endocrine system. The manifestation of such toxicity may be noted as alterations in sexual behavior, 

fertility, pregnancy outcomes, or modifications in other functions that are dependent on the integrity of 

this system. 

Retrospective Study—A type of cohort study based on a group of persons known to have been exposed 

at some time in the past. Data are collected from routinely recorded events, up to the time the study is 

undertaken. Retrospective studies are limited to causal factors that can be ascertained from existing 

records and/or examining survivors of the cohort. 
PBBs and PBDEs 564 

12. GLOSSARY 

Risk—The possibility or chance that some adverse effect will result from a given exposure to a chemical. 

Risk Factor—An aspect of personal behavior or lifestyle, an environmental exposure, or an inborn or 

inherited characteristic that is associated with an increased occurrence of disease or other health-related 

event or condition. 

Risk Ratio—The ratio of the risk among persons with specific risk factors compared to the risk among 

persons without risk factors. A risk ratio greater than 1 indicates greater risk of disease in the exposed 

group compared to the unexposed. 

Short-Term Exposure Limit (STEL)—The American Conference of Governmental Industrial 

Hygienists (ACGIH) maximum concentration to which workers can be exposed for up to 15 minutes 

continually. No more than four excursions are allowed per day, and there must be at least 60 minutes 

between exposure periods. The daily Threshold Limit Value - Time Weighted Average (TLV-TWA) may 

not be exceeded. 

Standardized Mortality Ratio (SMR)—A ratio of the observed number of deaths and the expected 

number of deaths in a specific standard population. 

Target Organ Toxicity—This term covers a broad range of adverse effects on target organs or 

physiological systems (e.g., renal, cardiovascular) extending from those arising through a single limited 

exposure to those assumed over a lifetime of exposure to a chemical. 

Teratogen—A chemical that causes structural defects that affect the development of an organism. 

Threshold Limit Value (TLV)—An American Conference of Governmental Industrial Hygienists 

(ACGIH) concentration of a substance to which most workers can be exposed without adverse effect. 

The TLV may be expressed as a Time Weighted Average (TWA), as a Short-Term Exposure Limit 

(STEL), or as a ceiling limit (CL). 

Time-Weighted Average (TWA)—An allowable exposure concentration averaged over a normal 8-hour 

workday or 40-hour workweek. 

Toxic Dose(50) (TD50)—A calculated dose of a chemical, introduced by a route other than inhalation, 

which is expected to cause a specific toxic effect in 50% of a defined experimental animal population. 

Toxicokinetic—The study of the absorption, distribution, and elimination of toxic compounds in the 

living organism. 

Uncertainty Factor (UF)—A factor used in operationally deriving the Minimal Risk Level (MRL) or 

Reference Dose (RfD) or Reference Concentration (RfC) from experimental data. UFs are intended to 

account for (1) the variation in sensitivity among the members of the human population, (2) the 

uncertainty in extrapolating animal data to the case of human, (3) the uncertainty in extrapolating from 

data obtained in a study that is of less than lifetime exposure, and (4) the uncertainty in using lowestobserved-adverse-effect level (LOAEL) data rather than no-observed-adverse-effect level (NOAEL) data. 

A default for each individual UF is 10; if complete certainty in data exists, a value of 1 can be used; 

however, a reduced UF of 3 may be used on a case-by-case basis, 3 being the approximate logarithmic 

average of 10 and 1. 

Xenobiotic—Any chemical that is foreign to the biological system. 
PBBs and PBDEs A-1 

APPENDIX A. ATSDR MINIMAL RISK LEVELS AND WORKSHEETS 

The Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) [42 U.S.C. 

9601 et seq.], as amended by the Superfund Amendments and Reauthorization Act (SARA) [Pub. L. 99–

499], requires that the Agency for Toxic Substances and Disease Registry (ATSDR) develop jointly with 

the U.S. Environmental Protection Agency (EPA), in order of priority, a list of hazardous substances most 

commonly found at facilities on the CERCLA National Priorities List (NPL); prepare toxicological 

profiles for each substance included on the priority list of hazardous substances; and assure the initiation 

of a research program to fill identified data needs associated with the substances. 

The toxicological profiles include an examination, summary, and interpretation of available toxicological 

information and epidemiologic evaluations of a hazardous substance. During the development of 

toxicological profiles, Minimal Risk Levels (MRLs) are derived when reliable and sufficient data exist to 

identify the target organ(s) of effect or the most sensitive health effect(s) for a specific duration for a 

given route of exposure. An MRL is an estimate of the daily human exposure to a hazardous substance 

that is likely to be without appreciable risk of adverse noncancer health effects over a specified duration 

of exposure. MRLs are based on noncancer health effects only and are not based on a consideration of 

cancer effects. These substance-specific estimates, which are intended to serve as screening levels, are 

used by ATSDR health assessors to identify contaminants and potential health effects that may be of 

concern at hazardous waste sites. It is important to note that MRLs are not intended to define clean-up or 

action levels. 

MRLs are derived for hazardous substances using the no-observed-adverse-effect level/uncertainty factor 

approach. They are below levels that might cause adverse health effects in the people most sensitive to 

such chemical-induced effects. MRLs are derived for acute (1–14 days), intermediate (15–364 days), and 

chronic (365 days and longer) durations and for the oral and inhalation routes of exposure. Currently, 

MRLs for the dermal route of exposure are not derived because ATSDR has not yet identified a method 

suitable for this route of exposure. MRLs are generally based on the most sensitive chemical-induced end 

point considered to be of relevance to humans. Serious health effects (such as irreparable damage to the 

liver or kidneys, or birth defects) are not used as a basis for establishing MRLs. Exposure to a level 

above the MRL does not mean that adverse health effects will occur. 

MRLs are intended only to serve as a screening tool to help public health professionals decide where to 

look more closely. They may also be viewed as a mechanism to identify those hazardous waste sites that 
PBBs and PBDEs A-2 

APPENDIX A 

are not expected to cause adverse health effects. Most MRLs contain a degree of uncertainty because of 

the lack of precise toxicological information on the people who might be most sensitive (e.g., infants, 

elderly, nutritionally or immunologically compromised) to the effects of hazardous substances. ATSDR 

uses a conservative (i.e., protective) approach to address this uncertainty consistent with the public health 

principle of prevention. Although human data are preferred, MRLs often must be based on animal studies 

because relevant human studies are lacking. In the absence of evidence to the contrary, ATSDR assumes 

that humans are more sensitive to the effects of hazardous substance than animals and that certain persons 

may be particularly sensitive. Thus, the resulting MRL may be as much as 100-fold below levels that 

have been shown to be nontoxic in laboratory animals. 

Proposed MRLs undergo a rigorous review process: Health Effects/MRL Workgroup reviews within the 

Division of Toxicology, expert panel peer reviews, and agency-wide MRL Workgroup reviews, with 

participation from other federal agencies and comments from the public. They are subject to change as 

new information becomes available concomitant with updating the toxicological profiles. Thus, MRLs in 

the most recent toxicological profiles supersede previously published levels. For additional information 

regarding MRLs, please contact the Division of Toxicology, Agency for Toxic Substances and Disease 

Registry, 1600 Clifton Road NE, Mailstop F-32, Atlanta, Georgia 30333. 
PBBs and PBDEs A-3 

APPENDIX A 

MINIMAL RISK LEVEL (MRL) WORKSHEET 

Chemical name: Polybrominated Diphenyl Ethers (PBDEs) 

 [Lower Brominated Diphenyl Ethers] 

CAS number(s): 32534-81-9 (pentaBDE), 32536-52-0 (octaBDE) 

Date: September 2, 2004 

Profile status: Final Post Public Comment 

Route: [X] Inhalation [ ] Oral 

Duration: [ ] Acute [X] Intermediate [ ] Chronic 

Key to figure: 3 

Species: Rat 

Minimal Risk Level: 0.006 [ ] mg/kg/day [ ] ppm [X] mg/m3

Reference: Great Lakes Chemical Corporation. 2000. A 90-day inhalation toxicity study of 

octabromodiphenyl oxide in albino rats, dated 04/04/02. Submitted to the U.S. EPA under TCSA Section 

8E, Fiche no. OTS0574171-1. 

Experimental design: This is an unpublished industry-sponsored study in which a commercial octaBDE 

product (Lot No. 9525DA23B, bromine content 78.7%, composition and purity not otherwise specified) 

was administered to groups of 10 male and 10 female Crl:CD(SD)IGS BR rats, via nose-only inhalation 

as a dust aerosol, in measured concentrations of 0 (filtered air-only), 1.1, 16 or 202 mg/m3

 for 

6 hours/day, 5 days/week, for 13 weeks. The mean MMADs in the low to high level groups were 2.0, 

2.7, and 2.8 microns; the corresponding mean GSDs were 3.37, 3.72, and 3.01. Clinical and physical 

signs, body weight, food consumption, and survival were evaluated throughout the study. Ophthalmic, 

hematology (11 indices), serum chemistry (18 indices), and serum thyroid hormone (TSH, total T3, and 

total T4) evaluations were performed near the end of the exposure period. Urinalyses were not conducted. 

Comprehensive necropies, organ weight measurements, and histological examinations (including 

respiratory tract and thyroids) were performed following exposure termination. 

Effects noted in study and corresponding doses: Hepatic, nasal, lung, thyroid, and ovarian effects were 

observed. The liver was affected in both sexes as shown by dose-related increases in centrilobular 

hepatocellular hypertrophy at ≥16 mg/m3

 and liver weight (absolute and relative) at 202 mg/m3. Total 

incidences of centrilobular hepatocellular hypertrophy in the 0, 1.1, 16, and 202 mg/m3

 groups were 

1/10 (minimal), 0/10, 3/10 (all minimal), and 10/10 (6 minimal, 2 mild, 2 moderate) in males, and 0/10, 

0/10, 3/10 (all minimal), and 6/10 (3 minimal, 3 mild) in females. Changes in nasal Goblet cells were 

increased at 202 mg/m3

, but showed no clear dose-related increasing trends for incidence and severity. 

Total incidences of Goblet cell hypertrophy (minimal or mild) were slightly increased in nasal level II of 

both sexes at ≥1.1 mg/m3

; incidences in 0, 1.1, 16, and 202 mg/m3 exposure groups were 4/10 (all 

minimal), 9/10 (7 minimal, 2 mild), 6/10 (all minimal), and 10/10 (9 minimal, 1 mild) in males, and 

2/10 (all minimal), 6/10 (all minimal), 4/10 (all minimal), and 8/10 (all minimal) in females. Goblet cell 

hypertrophy was also slightly increased in nasal level IV in males at 202 mg/m3

 (4/10, 0/10, 1/10, and 

8/10, all minimal severity, not increased in females). Histological changes in the lungs included alveolar 

histiocytosis and chronic active inflammation that were only clearly induced at 202 mg/m3

. Total 

incidences of alveolar histiocytosis at 0, 1.1, 16, and 202 mg/m3

 were 3/10 (2 mild, 1 minimal), 5/10 (all 

minimal), 5/10 (all minimal), and 10/10 (5 minimal, 3 mild, 2 moderate) in males, and 0/10, 5/10 (all 

minimal), 2/10 (all minimal), and 10/10 (1 minimal, 7 mild, 2 moderate) in females. Corresponding total 

incidences of chronic active lung inflammation were 0/10, 0/10, 2/10 (both minimal), and 

10/10 (5 minimal, 4 mild, 1 moderate) in males, and 0/10, 1/10 (minimal), 1/10 (minimal), and 

10/10 (2 minimal, 5 mild, 3 moderate) in females. Gross lung changes also occurred in both sexes at 

202 mg/m3

; these included lung firmness and white discoloration and/or enlargement in the bronchial 
PBBs and PBDEs A-4 

APPENDIX A 

and/or mediastinal lymph nodes. The lymph node effects correlated with the histological finding of 

granulomatous inflammation. There were no exposure-related gross or histopathological changes in the 

spleen, bone marrow, thymus, or other tissues, including thyroid. Thyroid hormone assessments, 

however, showed exposure-related decreases in mean thyroxine (total T4) at ≥16 mg/m3

 in both sexes, and 

increases in thyroid stimulating hormone (TSH) at ≥16 mg/m3

 in males and 202 mg/m3 in females. The 

changes were usually statistically significant (p<0.05 or p<0.01) compared to controls and considered to 

be consistent with chemical-induced hypothyroidism. There were no serum T3 changes. Qualitative 

histological evaluations of step sections of ovaries showed an absence of corpora lutea in 3/10 females at 

202 mg/m3

, compared to 0/10 in the control and lower exposure groups. This 30% incidence was 

interpreted to be a treatment-related effect because an absence of corpora lutea was considered unusual in 

rats at 20 weeks of age. 

Other findings included some hematological alterations in 202 mg/m3

 females that were not considered to 

be exposure-related (slightly increased mean activated partial thromboplastin time, and decreased mean 

MCH and MCHC without effects on RBC counts, hematocrit, or hemoglobin levels). Serum chemistry 

evaluations showed that cholesterol was significantly increased (39.8% less than controls, p<0.01) in 

202 mg/m3

 females, but the magnitude of the elevation was not considered toxicologically significant. 

Some other statistically significant serum chemistry alterations (increased mean globulin and total 

protein, decreased albumin/globulin ratio) also occurred in the 202 mg/m3

 females, but were not 

considered exposure-related due to small magnitudes of changes and lack of similar findings in the males. 

Dose and end point used for MRL derivation: 1.1 mg/m3 

[X] NOAEL [ ] LOAEL 

Considering the unclear adversity of minimal severity Goblet cell hypertrophy, lack of clear dose-related 

increasing trends for incidence and severity of this nasal effect, identification of both a NOAEL 

(1.1 mg/m3

) and LOAEL (16 mg/m3) for changes in thyroid hormones, and abundant evidence for thyroid 

effects of PBDEs in oral studies, the NOAEL for effects on thyroid hormones is the most appropriate 

basis for derivation of the MRL. 

Uncertainty factors used in MRL derivation: 

[X] 3 for extrapolation from animals to humans with dosimetric adjustments 

[X] 10 for human variability 

[X] 3 modifying factor for incomplete data base 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? NA 

Was a conversion used from intermittent to continuous exposure? The NOAEL was adjusted to 

continuous exposure as follows: 1.1 mg/m3 x 6 hours/24 hours x 5 days/7 days = 0.196 mg/m3

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: The 

human equivalent NOAEL (NOAELHEC) was calculated from the duration-adjusted NOAEL (NOAELADJ) 

using EPA RfC methodology as follows: 

 NOAELHEC = NOAELADJ x RDDR = 0.196 mg/m3 x 2.7 = 0.53 mg/m3 

The regional deposited dose ratio (RDDR) for the extrathoracic (ET) region was used to extrapolate 

deposited doses in rats to deposited doses in humans. The following parameters were used to calculate 
PBBs and PBDEs A-5 

APPENDIX A 

the RDDR of 2.7: mean particle size (MMAD) of 2.0 μm with a mean GSD (sigma g) of 3.37; default 

human body weight of 70 kg, and a default female F344 rat body weight of 180 g. 

Based on these values, the MRL is derived as follows: 

 MRL = NOAELHEC ÷ (UF x MF) = 0.53 ÷ (30 x 3) = 0.006 mg/m3

 

Other additional studies or pertinent information that lend support to this MRL: This is the only 

intermediate-duration inhalation study of PBDEs. 

The thyroid is a sensitive target of lower brominated BDEs in orally exposed animals. A NOAEL for 

reduced serum T4 hormone levels in fetal rats that were exposed to pentaBDE (Zhou et al. 2002) was used 

as the basis for the acute oral MRL for lower brominated BDEs. This study is supported by findings of 

reduced serum T4 levels in weanling rats that were acutely exposed to octaBDE and pentaBDE as 

neonates (Zhou et al. 2001). Thyroid effects that mainly included reduced serum T4 hormone levels and 

follicular cell hyperplasia were observed in a number of other animal studies of acute- or intermediateduration oral exposure to penta- and octaBDEs (Fowles et al. 1994; Hallgren et al. 2001; WIL Research 

Laboratories 1984). 

Agency Contact (Chemical Manager): Dr. Hana Pohl 
PBBs and PBDEs A-6 

APPENDIX A 

MINIMAL RISK LEVEL (MRL) WORKSHEET 

Chemical name: Polybrominated Biphenyls (PBBs) 

CAS number(s): 36355-01-8 (unspecified hexabromo mixture) 

Date: September 2, 2004 

Profile status: Final Post Public Comment 

Route: [ ] Inhalation [X] Oral 

Duration: [X] Acute [ ] Intermediate [ ] Chronic 

Key to figure: 4 

Species: Rat 

Minimal Risk Level: 0.01 [X] mg/kg/day [ ] ppm [ ] mg/m3 

Reference: Allen-Rowlands CF, Castracane VD, Hamilton MG, et al. 1981. Effect of polybrominated 

biphenyls (PBB) on the pituitary - thyroid axis of the rat. Proc Soc Exp Biol Med 166:506-514. 

Experimental design: Groups of 8–11 male rats were treated with 0, 1, 3, or 6 mg/kg/day doses of an 

unspecified mixture of PBBs in lecithin liposomes by gavage for 10 days. Plasma T4 was assayed on 

treatment days 10 and 20. Other end points were evaluated on treatment day 20; these included plasma 

TSH levels, 5-hour thyroid uptake of 131I, incorporation of 131I into monoiodotyrosine, diiodotyrosine, T3

or T4, amount of intrathyroidal iodide, thyroid and liver weights, and body weights. Differences between 

mean values for the measured parameters in the control and PBB-treated groups were analyzed with the 

Student’s t-test, with a P value of 0.05 considered as statistically significant. 

Effects noted in study and corresponding doses: Plasma (T4) was significantly (p<0.05) decreased at 

≥3 mg/kg/day after 10 and 20 days; this reduction was both dose- and time-dependent. Plasma TSH 

levels were significantly elevated (p<0.01) at 6 mg/kg/day. The 6 mg/kg dose also produced a significant 

increase (p<0.01) in the 5-hour thyroid uptake of 131I and a significant depression (p<0.01) in the 

incorporation of 131I into monoiodotyrosine, without any apparent effect on the incorporation of 131I into 

diiodotyrosine, T3 or T4. There was a significant increase (p<0.01) of intrathyroidal iodide (nine rats/dose 

evaluated). At ≥3 mg/kg, the absolute thyroid weights were significantly increased (p<0.01) (not 

evaluated at 1 mg/kg). Relative liver weight was significantly increased at ≥1 mg/kg/day, but no 

treatment related effects on body weight were observed. The 1 mg/kg/day dose is considered a NOAEL. 

Despite the fact that the inappropriate statistical test (t-test, rather than an ANOVA with multiple 

comparison tests) was used to analyze these data, ATSDR is confident with the designation of the 

NOAEL and LOAEL values. The data in the manuscript are presented graphically, with animal numbers 

presented as a range (8–11 animals/group); thus, an ANOVA could not be performed from the published 

report. However, using the graphical data, the change in plasma T4 levels in the 3 mg/kg/day groups is 

clearly on the order of 20–30%, which represents a biologically significant change. As such, the 

identification of 3 mg/kg/day as a LOAEL, and 1 mg/kg/day as a NOAEL, is not contraindicated by the 

lack of appropriate statistical analysis. 

Dose and end point used for MRL derivation: 

[X] NOAEL [ ] LOAEL 

Uncertainty factors used in MRL derivation: 

[X] 10 for extrapolation from animals to humans 

[X] 10 for human variability 
PBBs and PBDEs A-7 

APPENDIX A 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? NA (gavage study) 

Was a conversion used from intermittent to continuous exposure? NA 

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: NA 

Other additional studies or pertinent information that lend support to this MRL: It is well documented in 

intermediate-duration studies that the thyroid is a target of PBBs showing a spectrum of effects, including 

decreases in serum T3 and T4 hormone, thyroid enlargement, effects in the follicular cells (e.g., reduced 

size, hyperplasia with columnar appearance, and papillary projections) and accumulation of colloid 

droplets (Akoso et al. 1982a, 1982b; Byrne et al. 1987; Gupta and Moore 1979; Kasza et al. 1978a; Norris 

et al. 1975a, 1975b, 1975c; NTP 1983; Sepkovic and Byrne 1984; Sleight et al. 1978). 

Agency Contact (Chemical Manager): Dr. Hana Pohl 
PBBs and PBDEs A-8 

APPENDIX A 

MINIMAL RISK LEVEL (MRL) WORKSHEET 

Chemical name: Polybrominated Diphenyl Ethers (PBDEs) 

 [Lower Brominated Diphenyl Ethers] 

CAS number(s): 32534-81-9 (pentaBDE), 32536-52-0 (octaBDE) 

Date: September 2, 2004 

Profile status: Final Post Public Comment 

Route: [ ] Inhalation [X] Oral 

Duration: [X] Acute [ ] Intermediate [ ] Chronic 

Key to figure: 27 

Species: Rat 

Minimal Risk Level: 0.03 [X] mg/kg/day [ ] ppm [ ] mg/m3

Reference: Zhou T, Taylor MM, DeVito MJ, et al. 2002. Developmental exposure to brominated 

diphenyl ethers results in thyroid hormone disruption. Toxicol Sci 66:105-116. 

Experimental design: Groups of 47, 47, 55, and 55 primparous Long-Evans rats were administered a 

technical pentaBDE mixture (DE-71) in corn oil by gavage from gestation day (GD) 6 through postnatal 

day (PND) 21, except for PND 0 (day of birth). DE-71 is a technical mixture consisting primarily of tetra 

and penta congeners. Dams were sacrificed on GD 20 and PND 22 and offspring were sacrificed on 

GD 20 and PNDs 4, 14, 36, and 90). Serum and liver samples were obtained from a minimum of eight 

litters at each age point and analyzed for thyroid hormone (total T4 and T3) concentrations, liver weight, 

and hepatic microsomal enzyme (EROD, PROD, UDPGT) activities. Serum T3 was not assayed in 

GD 20 fetuses due to insufficient serum sample volume. Other study end points included number of pups 

delivered on GD 21, number and sex of pups on PNDs 4, 7, 14, and 21, body weight of pups on PNDs 4, 

7, 14, 21, 36, and 90, eye opening status (pups with at least one eye open) on PNDs 11–18, and maternal 

body weight on GD 6 through PND 21. 

The litter was the statistical unit for all analyses. Analysis of variance (ANOVA) was used to analyze for 

effects of treatment and interactions. If there was more than one independent variable, significant 

interactions were followed by step-down ANOVA tests for each independent variable (e.g., treatment and 

age). A nested design was used when more than one reading for each litter was obtained (e.g., body 

weights for males and females from the same litter). Repeated-measure ANOVAs were applied to data 

on dam body weights, preweaning offspring body weights (PNDs 4–21), and eye opening. Postweaning 

offspring body weights (PNDs 36–91) were analyzed with a two-way ANOVA, with time and dose as 

independent variables and litter nested under treatment. For significant effects of treatment, Duncan’s 

Multiple Comparison test was used for mean contrast comparisons. The fetal T4 data were analyzed with 

the Kruskal-Wallis test followed by a Dunn Multiple Comparison tests (due to a lack of homogeneity of 

variance). A significance level of 0.05 was used for all statistical tests. Benchmark dose estimates were 

determined for alterations in thyroid hormones and hepatic enzyme activity using the U.S. EPA 

Benchmark Dose Software (BMDS, V 1.3). 

Effects noted in study and corresponding doses: No treatment-related effects on gestation length, litter 

size, sex ratio, viability index (percent of pups surviving until day 4), maternal or offspring body weights, 

or offspring eye opening were observed. The perinatal maternal exposure to PentaPBE caused significant 

(p<0.05) decreases in serum total T4 in dams at 30 mg/kg/day on GD 20 and PND 22 (48 and 44%, 

respectively, relative to controls), and in fetuses and offspring at ≥10 mg/kg/day on GD 20 (at least 15% 

reduced) and PNDs 4 and 14 (50 and 64% maximal in the 10 and 30 mg/kg/day groups, respectively). 

The effect on T4 concentrations in offspring was age-dependent as values returned to control levels by 

PND 36. There were no exposure-related effects on serum total T3 concentrations at any time in the dams 
PBBs and PBDEs A-9 

APPENDIX A 

or offspring. Relative liver weight was significantly (p<0.05) increased in dams from GD 20 to PND 22 

(approximately 8% above controls) at 30 mg/kg/day and offspring during the early preweaning period at 

≥10 mg/kg/day. Liver weights in offspring were maximal at 30 mg/kg/day on PNDs 4 and 14 (35 and 

39% above controls, respectively) and returned to control levels by PND 36. Microsomal enzyme activity 

was increased in both dams and offspring as shown by significantly elevated hepatic EROD and PROD at 

≥10 mg/kg/day in dams on GD 20 and PND 22 and offspring on GD 20 and PNDs 4, 14, and 36, and 

signficantly elevated UDPGT at 30 mg/kg/day in dams on GD 20 and PND 22 and offspring on GD 20 

and PNDs 4 and 14. Benchmark dose analysis of the data found that the BMD and BMDL (95% lower 

confidence bound confidence limit on the effective dose) resulting in a 20% reduction of serum T4

(LED20) were 2.36 and 0.94 mg/kg/day, respectively. The respective BMD and BMDL resulting in 50% 

increased enzyme activity were 0.43 and 0.31 mg/kg/day for EROD, 0.48 and 0.36 mg/kg/day for PROD, 

and 5.50 and 3.41 mg/kg/day for UDPGT. 

Dose and end point used for MRL derivation: 

[X] NOAEL [ ] LOAEL 

A NOAEL of 1 mg/kg/day and a LOAEL of 10 mg/kg/day were identified for reduced serum T4 levels in 

fetal rats on GD 20. 

Uncertainty factors used in MRL derivation: 

[X] 10 for extrapolation from animals to humans 

[X] 3 for human variability 

A component factor of 10 was not used for human variability because the MRL is based on effect levels 

identified in a sensitive subgroup (i.e., neonates exposed in utero). 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? NA (gavage study) 

Was a conversion used from intermittent to continuous exposure? NA 

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: NA 

Other additional studies or pertinent information that lend support to this MRL: The NOAEL of 

1 mg/kg/day and LOAEL of 10 mg/kg/day for reduced serum T4 hormone levels in fetal rats that were 

exposed to pentaBDE in the acute MRL study (Zhou et al. 2002) are supported by a NOAEL of 

3 mg/kg/day and LOAEL of 10 mg/kg/day for reduced serum T4 levels in weanling rats that were exposed 

to octaBDE for 4 days (Zhou et al. 2001). Thyroid hormone levels were determined in groups of eight 

weanling (28-day-old) female Long-Evans that were treated by gavage for 4 days with commercial 

mixtures of decaBDE (DE-83R) or octaBDE (DE-79) in doses of 0.3, 1, 3, 10, 30, 60, or 100 mg/kg/day, 

or pentaBDE (DE-71) in doses of 0.3, 1, 3, 10, 30, 100, or 300 mg/kg/day. The animals were sacrificed 

on the day after the last exposure and evaluated for changes in serum total T4, total T3, and TSH, hepatic 

microsomal EROD, PROD, and UDPGT activities, and body and liver weight. No dose-related effects on 

any of the measured parameters were observed for decaBDE. OctaBDE and pentaBDE caused reduced 

thyroid hormone levels and increased microsomal enzyme activities. 

Effects of octaBDE included dose-related reductions in serum total T4 levels with statistically significant 

(p<0.05) decreases occurring at ≥10 mg/kg/day and a 70% maximum decrease compared to controls at 

100 mg/kg/day (Zhou et al. 2001). Serum T3 levels were reduced less than T4 levels and were significant 

at ≥60 mg/kg/day with a maximum reduction of 25% at the highest dose of 100 mg/kg/day. There were 
PBBs and PBDEs A-10 

APPENDIX A 

no exposure-related changes in serum total TSH concentrations. Hepatic microsomal EROD and UDPGT 

activities and relative liver weight were significantly (p<0.05) increased at ≥30 mg/kg/day and PROD 

activity was increased at ≥10 mg/kg/day in dose-related responses. No treatment-related effects on body 

weight or visible signs of toxicity were observed. Benchmark dose analysis of the octaBDE data found 

that the BMD and BMDL resulting in a 20% reduction in thyroid hormones (LED20) were 9.25 and 

5.29 mg/kg/day, respectively, for serum T4 and 53.38 and 11.98 mg/kg/day, respectively, for serum T3. 

The respective BMD and BMDL resulting in 50% increased enzyme activity were 3.66 and 

2.45 mg/kg/day for EROD, 0.53 and 0.40 mg/kg/day for PROD, and 21.17 and 11.03 mg/kg/day for 

UDPGT. 

Effects of pentaBDE included dose-related reductions in serum total T4 levels with significant (p<0.05) 

decreases occurring at ≥30 mg/kg/day and an 80% maximum decrease compared to controls 

300 mg/kg/day (Zhou et al. 2001). Serum total T3 levels were reduced less than T4 levels and were 

significant at ≥100 mg/kg/day with a maximum reduction of 30% at the highest dose of 300 mg/kg/day. 

There were no exposure-related changes in serum total TSH concentrations. Hepatic microsomal EROD 

and PROD activities and relative liver weight were significantly (p<0.05) increased at ≥10 mg/kg/day and 

UDPGT activity was increased at ≥30 mg/kg/day in dose-related responses. No treatment-related effects 

on body weight or visible signs of toxicity were observed. The BMD and BMDL resulting in a 20% 

reduction in thyroid hormones (LED20) were 12.74 and 6.95 mg/kg/day, respectively, for serum T4 and 

32.94 and 8.56, mg/kg/day, respectively, for serum T3. The respective BMD and BMDL resulting in 50% 

increased enzyme activity were 2.88 and 1.82 mg/kg/day for EROD, 0.81 and 0.54 mg/kg/day for PROD, 

and 9.51 and 5.83 mg/kg/day for UDPGT. 

Thyroid effects that mainly included reduced serum T4 hormone levels and follicular cell hyperplasia 

have been observed in a number of other studies of PBDEs in orally-exposed animals. Other acuteduration studies showed decreases in serum T4 in rats and mice exposed to ≥18 mg/kg/day pentaBDE for 

14 days (Fowles et al. 1994; Hallgren et al. 2001). Effects observed in intermediate-duration studies 

include thyroid hyperplasia in rats exposed to ≥8 mg/kg/day octaBDE for 30 days (Norris et al. 1973, 

1975b) and reduced serum T4 in rats exposed to ≥10 mg/kg/day pentaBDE for 90 days (WIL Research 

Laboratories 1984). Intermediate-duration exposure to a 77% decaBDE/22% nonaBDE commercial 

mixture caused thyroid hyperplasia in rats at doses ≥80 mg/kg/day for 30 days (Norris et al. 1973, 1975b). 

Chronic (103-week) exposure to high-purity decaBDE (≥97%) did not induce thyroid histopathological 

changes in rats at ≤2,550 mg/kg/day, although follicular cell hyperplasia developed in mice exposed to 

2,240 mg/kg/day (NTP 1986). 

Agency Contact (Chemical Manager): Dr. Hana Pohl 
PBBs and PBDEs A-11 

APPENDIX A 

MINIMAL RISK LEVEL (MRL) WORKSHEET 

Chemical name: Polybrominated Diphenyl Ethers (PBDEs) 

 [Lower Brominated Diphenyl Ethers] 

CAS number(s): 32534-81-9 (pentaBDE), 32536-52-0 (octaBDE) 

Date: September 2, 2004 

Profile status: Final Post Public Comment 

Route: [ ] Inhalation [X] Oral 

Duration: [ ] Acute [X] Intermediate [ ] Chronic 

Key to figure: 36 

Species: Rat 

Minimal Risk Level: 0.007 [X] mg/kg/day [ ] ppm [ ] mg/m3

Reference: WIL Research Laboratories. 1984. 90-day dietary study in rats with pentabromodiphenyl 

oxide (DE-7) (Volume I-II). Submitted to U.S. EPA under TSCA Section 8D, Fiche no. OTS0524336. 

Experimental design: PentaBDE (commercial mixture DE-71) was administered to Sprague-Dawley rats 

in the diet at dosage levels of 0, 2, 10, or 100 mg/kg/day for 90 days. Ten rats/sex/group were sacrificed 

after 4 weeks, 10 rats/sex/group were sacrificed at the end of the 90-day exposure period, and the 

remaining rats were sacrificed after a 6-week recovery period (5 rats/sex/group) or 24-week recovery 

period (5 rats/sex/group). Animals were observed daily for general appearance, behavior, signs of overt 

toxicity, and mortality during the dosing and recovery periods. Body weight and food consumption were 

measured during the dosing period and the first 4 weeks of the recovery period. Hematology, clinical 

chemistry (including serum T3 and T4), and urinalysis parameters, urine and liver porphyrin levels, and 

serum bromide levels were evaluated in 10 rats/sex/group at weeks 4 and 13. Bromine levels in liver, 

lung, kidney, thymus, and thyroid were evaluated at weeks 4, 13, 19, and 37. Gross necropsies and organ 

weight measurements (brain, gonads, heart, liver, kidneys, thymus, and thyroid) were performed on all 

rats. Histological examinations included liver, lung, kidney, thymus, thyroid, and gross lesions in all dose 

groups at the week 4, 13, 19, and 37 sacrifices and in all tissues (comprehensive evaluation) in the 0 and 

100 mg/kg/day groups. 

All statistical analyses were conducted using two-tailed tests for a minimum significance level of 5% 

comparing treatment groups to the controls. Analysis of weekly body weights, body weight changes, 

food consumption, clinical laboratory values, and absolute and relative organ weights were analyzed by a 

one-way analysis of variance and Dunnett’s Test. The one-tailed Kolmogorov-Smirnov test was used for 

the 4- and 13-week histopathological diagnoses. 

Effects noted in study and corresponding doses: Effects observed were observed at ≥2 mg/kg/day and 

included histological changes in the liver after 4 and 13 weeks of exposure as well as increased bromine 

levels in essentially all measured tissues. Hepatocytomegaly occurred in males at 2 mg/kg/day and both 

sexes were affected at the higher doses. Incidences of hepatomegaly in the control to high dose groups at 

13 weeks were 0/10, 7/10, 10/10, and 10/10 in males and 0/10, 0/10, 8/10, and 10/10 (statistical tests not 

conducted) in females. Some hepatocytes in affected areas had vacuoles, which were empty in tissue 

sections and reportedly likely contained neutral lipid. Incidences of hepatocyte vacuolation at 13 weeks 

were 2/10, 4/10, 3/10, and 2/10 in males and 3/10, 5/10, 5/10, and 6/10 in females. The 

hepatocytomegaly was similar in incidence and severity after 4 and 13 weeks, appeared to be dose-related 

with respect to severity, and was not completely reversible as it was still observed in males at 

≥10 mg/kg/day and in females at 100 mg/kg/day at 24 weeks postexposure in lessened severity and 

incidence. Females in the 2 and 100 mg/kg/day groups appeared to have an increased incidence of 

degeneration and necrosis of individual liver parenchymal cells at 24 weeks postexposure; the 
PBBs and PBDEs A-12 

APPENDIX A 

investigators concluded that this condition may represent the final loss of previously damaged cells and 

probably should be considered compound-related. Incidences of individual liver cell degeneration/ 

necrosis in the female control to high dose groups at 24 weeks postexposure were 2/5, 5/5, 2/5, and 5/5. 

Small increases in thyroid hyperplasia appeared to occur in lower dose groups; incidences in the control 

to high dose groups at 13 weeks were 0/10, 2/10, 2/10, and 5/10 in males and 0/10, 0/10, 1/10, and 4/10 in 

females. The thyroid hyperplasia was very slight in severity at 2 and 10 mg/kg/day, very slight to slight 

severity at 100 mg/kg/day, and reversible in that it was no longer observed at 24 weeks postexposure in 

any animals; the thyroid changes therefore were mild and transient. Serum T4 levels were significantly 

reduced in both sexes at ≥10 mg/kg/day, indicating that 10 mg/kg/day is the LOAEL for thyroid effects. 

The slight thyroid hyperplasia and reductions in plasma T4 levels are likely indirect consequences of 

hepatic enzyme induction. No compound-related changes were observed in any tissues other than the 

liver and thyroid at any dose level. Other effects observed at ≥10 mg/kg/day included increased serum 

bromide levels in both sexes at 4 weeks (only increased in both sexes at 100 mg/kg/day at 13 weeks), and 

increased urine porphyrins in both sexes and liver porphyrins in females (liver porphyrins increased in 

males at 100 mg/kg/day). Effects observed at 100 mg/kg/day included decreased body weight gain in 

both sexes, decreased food consumption in females, and increased serum cholesterol in both sexes at 

weeks 4 and 13, and increased absolute and relative liver weights in both sexes at 13 weeks (returned to 

normal ranges after 24 weeks of recovery). 

Dose and end point used for MRL derivation: 

[ ] NOAEL [X] LOAEL 

The lowest tested dose, 2 mg/kg/day, is a minimal LOAEL for hepatic effects (hypertrophy, mild 

degeneration, and slight necrosis). 

Uncertainty factors used in MRL derivation: 

[X] 3 for extrapolation from a minimal LOAEL to a NOAEL 

[X] 10 for extrapolation from animals to humans 

[X] 10 for human variability 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? No (reported doses) 

Was a conversion used from intermittent to continuous exposure? NA 

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: NA 

Other additional studies or pertinent information that lend support to this MRL: The hepatotoxic potential 

of PBDE mixtures is well-documented in animals by oral exposure. The spectrum of observed hepatic 

effects includes microsomal enzyme induction, liver enlargement, and degenerative histopathologic 

alterations that progress to tumors. Repeated dietary exposure to PBDEs typically caused liver 

enlargement with or without degenerative changes, and effects were generally dose-related in incidence 

and severity, more frequent and pronounced in males than females, and more severe with octaBDE and 

pentaBDE than decaBDE. Hepatic effects induced by chronic exposure to decaBDE included 

degeneration and thrombosis in rats exposed to 2,240 mg/kg/day and centrilobular hypertrophy and 

granulomas in mice exposed to ≥3,200 mg/kg/day (NTP 1986). Data from other intermediate-duration 

studies that support selection of the 2 mg/kg/day critical LOAEL include hepatic LOAELs of 

5 mg/kg/day for cytomegaly (with vacuolation and necrosis at higher doses) in rats exposed to octaBDE 

for 13 weeks (IRDC 1977), 8 mg/kg/day for hepatocellular enlargement and vacuolation in rats exposed 
PBBs and PBDEs A-13 

APPENDIX A 

to octaBDE for 30 days (Norris et al. 1973, 1975b), and 9 mg/kg/day for hepatocellular enlargement and 

increased liver weight in rats exposed to octaBDE or pentaBDE for 28 days (IRDC 1976). 

Agency Contact (Chemical Manager): Dr. Hana Pohl 
PBBs and PBDEs A-14 

APPENDIX A 

MINIMAL RISK LEVEL (MRL) WORKSHEET 

Chemical name: Polybrominated Diphenyl Ethers (PBDEs) 

 [Decabromodiphenyl Ether (DecaBDPE)] 

CAS number(s): 1163-19-5 

Date: September 2, 2004 

Profile status: Final Post Public Comment 

Route: [ ] Inhalation [X] Oral 

Duration: [ ] Acute [X] Intermediate [ ] Chronic 

Key to figure: 13 

Species: Mouse 

Minimal Risk Level: 10 [X] mg/kg/day [ ] ppm [ ] mg/m3

Reference: Hardy ML, Schroeder R, Biesemeier J, et al. 2002. Prenatal oral (gavage) developmental 

toxicity study of decabromodiphenyl oxide in rats. Int J Toxicol 21:83-91. 

Experimental design: A commercial decaBDPE product (97.34% DBDPO, 2.66% nonaBDE/octaBDE) 

was administered to groups of 25 mated female Sprague-Dawley rats by gavage in corn oil in daily doses 

of 0, 100, 300, or 1,000 mg/kg/day on gestation days 0 through 19. Each female was sacrificed on 

gestation day 20 and necropsied. End points examined included maternal clinical observations, maternal 

body weight/weight gain and food consumption, maternal gravid uterine and liver weights, maternal gross 

lesions, total number of corpora lutea, uterine implantations, early and late resorptions, viable and 

nonviable fetuses, and fetal weight and sex. Fetuses were examined grossly (all fetuses), evaluated for 

skeletal/cartilaginous malformations and ossification variations (approximately half of each litter), and 

evaluated for visceral malformations (remaining fetuses). The experiment was designed to meet current 

(as of 2002) EPA/OPPTS and OECD guidelines for a prenatal developmental toxicity study. 

Effects noted in study and corresponding doses: No effects on any maternal end points (e.g., clinical 

signs, body weight, pregnancy rate, implantation, liver weight, necropsy findings) or fetal endpoints (e.g., 

body weight, sex ratio, or external, visceral or skeletal malformations) were observed in any dose group. 

Dose and end point used for MRL derivation: 1,000 mg/kg/day 

[X] NOAEL [ ] LOAEL 

No exposure-related maternal or developmental toxicity was found, indicating that the critical NOAEL is 

1,000 mg/kg/day, the highest tested dose. 

Uncertainty factors used in MRL derivation: 

[X] 10 for extrapolation from animals to humans 

[X] 10 for human variability 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? NA 

Was a conversion used from intermittent to continuous exposure? NA 

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: NA 
PBBs and PBDEs A-15 

APPENDIX A 

Other additional studies or pertinent information that lend support to this MRL: One intermediateduration oral systemic toxicity study of a high purity decaBDE commercial product was conducted. A 

commercial decaBDE mixture (94B97% pure) was fed to F344 rats (10/sex/level) and B6C3F1 mice 

(10/sex/level) in dietary concentrations of 0, 3,100, 6,300, 12,500, 25,000, or 50,000 ppm for 13 weeks 

(NTP 1986). Corresponding estimated daily doses are 0, 496, 992, 2,000, 4,000, or 8,000 mg/kg/day in 

rats, and 0, 589, 1,178, 2,375, 4,750, or 9,500 mg/kg/day in mice. The daily doses (mg/kg body 

weight/day) were estimated by multiplying the dietary concentrations (ppm; mg/kg food) by food factors 

of 0.16 kg food/kg body weight/day for rats and 0.19 kg food/kg body weight/day for mice. The food 

factors are based on the averages of male and female subchronic reference values for food consumption 

and body weight in F344 rats and B6C3F1 mice (EPA 1988a). All animals were observed daily, weighed 

weekly, and necropsied at the end of the exposure period. Comprehensive histological examinations were 

also performed at the end of the study, but were limited to the control and high-dose groups. No 

compound-related clinical signs, deaths, body weight or food consumption changes, gross pathology, or 

histopathology were observed in either species at any level of exposure. 

Based on the findings of the NTP (1986) study, the intermediate-duration NOAELs for systemic toxicity 

are 8,000 mg/kg/day in rats and 9,500 mg/kg/day in mice. The NOAEL for developmental toxicity is 

1,000 mg/kg/day (Hardy et al. 2002). Because doses higher than 1,000 mg/kg/day have not been tested 

for developmental toxicity, and the NTP (1986) study indicates that this dose is also a NOAEL for 

systemic toxicity, the 1,000 mg/kg/day developmental toxicity NOAEL is used as the basis for the MRL. 

Agency Contact (Chemical Manager): Dr. Hana Pohl 

PBBs and PBDEs B-1 

APPENDIX B. USER'S GUIDE 

Chapters 1 and 2 

Public Health Statement 

These chapters of the profile are health effects summaries written in non-technical language. Their 

intended audience is the general public, especially people living in the vicinity of a hazardous waste site 

or chemical release. If the Public Health Statements were removed from the rest of the document, it 

would still communicate to the lay public essential information about the chemical. 

The major headings in the Public Health Statements are useful to find specific topics of concern. The 

topics are written in a question and answer format. The answer to each question includes a sentence that 

will direct the reader to chapters in the profile that will provide more information on the given topic. 

Chapters 3 and 4 

Relevance to Public Health 

These chapters provide health effects summaries based on evaluations of existing toxicologic, 

epidemiologic, and toxicokinetic information. These summaries are designed to present interpretive, 

weight-of-evidence discussions for human health end points by addressing the following questions. 

 1. What effects are known to occur in humans? 

 2. What effects observed in animals are likely to be of concern to humans? 

 3. What exposure conditions are likely to be of concern to humans, especially around hazardous 

waste sites? 

The chapters cover end points in the same order that they appear within the Discussion of Health Effects 

by Route of Exposure section, by route (inhalation, oral, and dermal) and within route by effect. Human 

data are presented first, then animal data. Both are organized by duration (acute, intermediate, chronic). 

In vitro data and data from parenteral routes (intramuscular, intravenous, subcutaneous, etc.) are also 

considered in this chapter. 

The carcinogenic potential of the profiled substance is qualitatively evaluated, when appropriate, using 

existing toxicokinetic, genotoxic, and carcinogenic data. ATSDR does not currently assess cancer 

potency or perform cancer risk assessments. Minimal Risk Levels (MRLs) for noncancer end points (if 

derived) and the end points from which they were derived are indicated and discussed. 

Limitations to existing scientific literature that prevent a satisfactory evaluation of the relevance to public 

health are identified in the Chapter 5 Data Needs section. 

Interpretation of Minimal Risk Levels 

Where sufficient toxicologic information is available, ATSDR has derived MRLs for inhalation and oral 

routes of entry at each duration of exposure (acute, intermediate, and chronic). These MRLs are not 

meant to support regulatory action, but to acquaint health professionals with exposure levels at which 

adverse health effects are not expected to occur in humans. 
PBBs and PBDEs B-2 

APPENDIX B 

MRLs should help physicians and public health officials determine the safety of a community living near 

a chemical emission, given the concentration of a contaminant in air or the estimated daily dose in water. 

MRLs are based largely on toxicological studies in animals and on reports of human occupational 

exposure. 

MRL users should be familiar with the toxicologic information on which the number is based. Chapters 3 

and 4, "Relevance to Public Health," contain basic information known about the substance. Other 

sections such as Chapter 5 Section 5.9, "Interactions with Other Substances,” and Section 5.10, 

"Populations that are Unusually Susceptible" provide important supplemental information. 

MRL users should also understand the MRL derivation methodology. MRLs are derived using a 

modified version of the risk assessment methodology that the Environmental Protection Agency (EPA) 

provides (Barnes and Dourson 1988) to determine reference doses (RfDs) for lifetime exposure. 

To derive an MRL, ATSDR generally selects the most sensitive end point which, in its best judgement, 

represents the most sensitive human health effect for a given exposure route and duration. ATSDR 

cannot make this judgement or derive an MRL unless information (quantitative or qualitative) is available 

for all potential systemic, neurological, and developmental effects. If this information and reliable 

quantitative data on the chosen end point are available, ATSDR derives an MRL using the most sensitive 

species (when information from multiple species is available) with the highest no-observed-adverse-effect 

level (NOAEL) that does not exceed any adverse effect levels. When a NOAEL is not available, a 

lowest-observed-adverse-effect level (LOAEL) can be used to derive an MRL, and an uncertainty factor 

(UF) of 10 must be employed. Additional uncertainty factors of 10 must be used both for human 

variability to protect sensitive subpopulations (people who are most susceptible to the health effects 

caused by the substance) and for interspecies variability (extrapolation from animals to humans). In 

deriving an MRL, these individual uncertainty factors are multiplied together. The product is then 

divided into the inhalation concentration or oral dosage selected from the study. Uncertainty factors used 

in developing a substance-specific MRL are provided in the footnotes of the levels of significant exposure 

(LSE) Tables. 

Chapter 5 

Health Effects 

Tables and Figures for Levels of Significant Exposure (LSE) 

Tables and figures are used to summarize health effects and illustrate graphically levels of exposure 

associated with those effects. These levels cover health effects observed at increasing dose 

concentrations and durations, differences in response by species, MRLs to humans for noncancer end 

points, and EPA's estimated range associated with an upper- bound individual lifetime cancer risk of 1 in 

10,000 to 1 in 10,000,000. Use the LSE tables and figures for a quick review of the health effects and to 

locate data for a specific exposure scenario. The LSE tables and figures should always be used in 

conjunction with the text. All entries in these tables and figures represent studies that provide reliable, 

quantitative estimates of NOAELs, LOAELs, or Cancer Effect Levels (CELs). 

The legends presented below demonstrate the application of these tables and figures. Representative 

examples of LSE Table 5-1 and Figure 5-1 are shown. The numbers in the left column of the legends 

correspond to the numbers in the example table and figure. 
PBBs and PBDEs B-3 

APPENDIX B 

LEGEND 

See Sample LSE Table 3-1 (page B-6) 

(1) Route of Exposure. One of the first considerations when reviewing the toxicity of a substance 

using these tables and figures should be the relevant and appropriate route of exposure. Typically 

when sufficient data exists, three LSE tables and two LSE figures are presented in the document. 

The three LSE tables present data on the three principal routes of exposure, i.e., inhalation, oral, 

and dermal (LSE Table 5-1, 5-2, and 5-3, respectively). LSE figures are limited to the inhalation 

(LSE Figure 5-1) and oral (LSE Figure 5-2) routes. Not all substances will have data on each 

route of exposure and will not, therefore, have all five of the tables and figures. 

(2) Exposure Period. Three exposure periods—acute (less than 15 days), intermediate (15–

364 days), and chronic (365 days or more)—are presented within each relevant route of exposure. 

In this example, an inhalation study of intermediate exposure duration is reported. For quick 

reference to health effects occurring from a known length of exposure, locate the applicable 

exposure period within the LSE table and figure. 

(3) Health Effect. The major categories of health effects included in LSE tables and figures are 

death, systemic, immunological, neurological, developmental, reproductive, and cancer. 

NOAELs and LOAELs can be reported in the tables and figures for all effects but cancer. 

Systemic effects are further defined in the "System" column of the LSE table (see key number 

18). 

(4) Key to Figure. Each key number in the LSE table links study information to one or more data 

points using the same key number in the corresponding LSE figure. In this example, the study 

represented by key number 18 has been used to derive a NOAEL and a Less Serious LOAEL 

(also see the two "18r" data points in sample Figure 5-1). 

(5) Species. The test species, whether animal or human, are identified in this column. Chapters 3 

and 4, "Relevance to Public Health," cover the relevance of animal data to human toxicity and 

Section 5.4, "Toxicokinetics," contains any available information on comparative toxicokinetics. 

Although NOAELs and LOAELs are species specific, the levels are extrapolated to equivalent 

human doses to derive an MRL. 

(6) Exposure Frequency/Duration. The duration of the study and the weekly and daily exposure 

regimen are provided in this column. This permits comparison of NOAELs and LOAELs from 

different studies. In this case (key number 18), rats were exposed to “Chemical x” via inhalation 

for 6 hours/day, 5 days/week, for 13 weeks. For a more complete review of the dosing regimen 

refer to the appropriate sections of the text or the original reference paper (i.e., Nitschke et al. 

1981). 

(7) System. This column further defines the systemic effects. These systems include respiratory, 

cardiovascular, gastrointestinal, hematological, musculoskeletal, hepatic, renal, and 

dermal/ocular. "Other" refers to any systemic effect (e.g., a decrease in body weight) not covered 

in these systems. In the example of key number 18, one systemic effect (respiratory) was 

investigated. 

(8) NOAEL. A NOAEL is the highest exposure level at which no harmful effects were seen in the 

organ system studied. Key number 18 reports a NOAEL of 3 ppm for the respiratory system, 

which was used to derive an intermediate exposure, inhalation MRL of 0.005 ppm (see 

footnote "b"). 
PBBs and PBDEs B-4 

APPENDIX B 

(9) LOAEL. A LOAEL is the lowest dose used in the study that caused a harmful health effect. 

LOAELs have been classified into "Less Serious" and "Serious" effects. These distinctions help 

readers identify the levels of exposure at which adverse health effects first appear and the 

gradation of effects with increasing dose. A brief description of the specific end point used to 

quantify the adverse effect accompanies the LOAEL. The respiratory effect reported in key 

number 18 (hyperplasia) is a Less Serious LOAEL of 10 ppm. MRLs are not derived from 

Serious LOAELs. 

(10) Reference. The complete reference citation is given in Chapter 11 of the profile. 

(11) CEL. A CEL is the lowest exposure level associated with the onset of carcinogenesis in 

experimental or epidemiologic studies. CELs are always considered serious effects. The LSE 

tables and figures do not contain NOAELs for cancer, but the text may report doses not causing 

measurable cancer increases. 

(12) Footnotes. Explanations of abbreviations or reference notes for data in the LSE tables are found 

in the footnotes. Footnote "b" indicates that the NOAEL of 3 ppm in key number 18 was used to 

derive an MRL of 0.005 ppm. 

LEGEND 

See Sample Figure 5-1 (page B-7) 

LSE figures graphically illustrate the data presented in the corresponding LSE tables. Figures help the 

reader quickly compare health effects according to exposure concentrations for particular exposure 

periods. 

(13) Exposure Period. The same exposure periods appear as in the LSE table. In this example, health 

effects observed within the acute and intermediate exposure periods are illustrated. 

(14) Health Effect. These are the categories of health effects for which reliable quantitative data 

exists. The same health effects appear in the LSE table. 

(15) Levels of Exposure. Concentrations or doses for each health effect in the LSE tables are 

graphically displayed in the LSE figures. Exposure concentration or dose is measured on the log 

scale "y" axis. Inhalation exposure is reported in mg/m3

 or ppm and oral exposure is reported in 

mg/kg/day. 

(16) NOAEL. In this example, the open circle designated 18r identifies a NOAEL critical end point in 

the rat upon which an intermediate inhalation exposure MRL is based. The key number 18 

corresponds to the entry in the LSE table. The dashed descending arrow indicates the 

extrapolation from the exposure level of 3 ppm (see entry 18 in the Table) to the MRL of 

0.005 ppm (see footnote "b" in the LSE table). 

(17) CEL. Key number 38r is one of three studies for which CELs were derived. The diamond 

symbol refers to a CEL for the test species-mouse. The number 38 corresponds to the entry in the 

LSE table. 
PBBs and PBDEs B-5 

APPENDIX B 

(18) Estimated Upper-Bound Human Cancer Risk Levels. This is the range associated with the upperbound for lifetime cancer risk of 1 in 10,000 to 1 in 10,000,000. These risk levels are derived 

from the EPA's Human Health Assessment Group's upper-bound estimates of the slope of the 

cancer dose response curve at low dose levels (q1*). 

(19) Key to LSE Figure. The Key explains the abbreviations and symbols used in the figure. 
Serious (ppm) (CEL, multiple organs) (CEL, lung tumors, nasal tumors) (CEL, lung tumors, hemangiosarcomas) 

11 

↓

20 

10 10 

LOAEL (effect) 

Less serious (ppm) 

9 

↓

10 (hyperplasia) 

SAMPLE 

NOAEL (ppm) 8 ↓3b 

7 

System 

↓

 Resp 

6 

Exposure frequency/ duration 

↓

13 wk 5 d/wk 6 hr/d 18 mo 5 d/wk 7 hr/d 89-104 wk 5 d/wk 6 hr/d 79-103 wk 5 d/wk 6 hr/d 

5 

Species 

↓

Rat Rat Rat Mouse 

TABLE 5-1. Levels of Significant Exposure to [Chemical x] - Inhalation 

Key to figurea

INTERMEDIATE EXPOSURE 

Systemic 18 

CHRONIC EXPOSURE 

Cancer 38 39 40 

a The number corresponds to entries in Figure 5-1. b Used to derive an intermediate inhalation Minimal Risk Level (MRL) of 5x10-3 ppm; dose adjusted for intermittent expo

 by an uncertainty factor of 100 (10 for extrapolation from animal to humans, 10 for human variability).

→→→→→

1 2 3 4 

12 


PBBs and PBDEs C-1 

APPENDIX C. ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

ACGIH American Conference of Governmental Industrial Hygienists 

ACOEM American College of Occupational and Environmental Medicine 

ADI acceptable daily intake 

ADME absorption, distribution, metabolism, and excretion 

AED atomic emission detection 

AFID alkali flame ionization detector 

AFOSH Air Force Office of Safety and Health 

ALT alanine aminotransferase 

AML acute myeloid leukemia 

AOAC Association of Official Analytical Chemists 

AOEC Association of Occupational and Environmental Clinics 

AP alkaline phosphatase 

APHA American Public Health Association 

AST aspartate aminotransferase 

atm atmosphere 

ATSDR Agency for Toxic Substances and Disease Registry 

AWQC Ambient Water Quality Criteria 

BAT best available technology 

BCF bioconcentration factor 

BEI Biological Exposure Index 

BMD benchmark dose 

BMR benchmark response 

BSC Board of Scientific Counselors 

C centigrade 

CAA Clean Air Act 

CAG Cancer Assessment Group of the U.S. Environmental Protection Agency 

CAS Chemical Abstract Services 

CDC Centers for Disease Control and Prevention 

CEL cancer effect level 

CELDS Computer-Environmental Legislative Data System 

CERCLA Comprehensive Environmental Response, Compensation, and Liability Act 

CFR Code of Federal Regulations 

Ci curie 

CI confidence interval 

CL ceiling limit value 

CLP Contract Laboratory Program 

cm centimeter 

CML chronic myeloid leukemia 

CPSC Consumer Products Safety Commission 

CWA Clean Water Act 

DHEW Department of Health, Education, and Welfare 

DHHS Department of Health and Human Services 

DNA deoxyribonucleic acid 

DOD Department of Defense 

DOE Department of Energy 

DOL Department of Labor 

DOT Department of Transportation 
PBBs and PBDEs C-2 

APPENDIX C 

DOT/UN/ Department of Transportation/United Nations/ 

 NA/IMCO North America/International Maritime Dangerous Goods Code 

DWEL drinking water exposure level 

ECD electron capture detection 

ECG/EKG electrocardiogram 

EEG electroencephalogram 

EEGL Emergency Exposure Guidance Level 

EPA Environmental Protection Agency 

F Fahrenheit 

F1 first-filial generation 

FAO Food and Agricultural Organization of the United Nations 

FDA Food and Drug Administration 

FEMA Federal Emergency Management Agency 

FIFRA Federal Insecticide, Fungicide, and Rodenticide Act 

FPD flame photometric detection 

fpm feet per minute 

FR Federal Register 

FSH follicle stimulating hormone 

g gram 

GC gas chromatography 

gd gestational day 

GLC gas liquid chromatography 

GPC gel permeation chromatography 

HPLC high-performance liquid chromatography 

HRGC high resolution gas chromatography 

HSDB Hazardous Substance Data Bank 

IARC International Agency for Research on Cancer 

IDLH immediately dangerous to life and health 

ILO International Labor Organization 

IRIS Integrated Risk Information System 

Kd adsorption ratio 

kg kilogram 

kkg metric ton 

Koc organic carbon partition coefficient 

Kow octanol-water partition coefficient 

L liter 

LC liquid chromatography 

LC50 lethal concentration, 50% kill 

LCLo lethal concentration, low 

LD50 lethal dose, 50% kill 

LDLo lethal dose, low 

LDH lactic dehydrogenase 

LH luteinizing hormone 

LOAEL lowest-observed-adverse-effect level 

LSE Levels of Significant Exposure 

LT50 lethal time, 50% kill 

m meter 

MA trans,trans-muconic acid 

MAL maximum allowable level 

mCi millicurie 

MCL maximum contaminant level 
PBBs and PBDEs C-3 

APPENDIX C 

MCLG maximum contaminant level goal 

MF modifying factor 

MFO mixed function oxidase 

mg milligram 

mL milliliter 

mm millimeter 

mmHg millimeters of mercury 

mmol millimole 

mppcf millions of particles per cubic foot 

MRL Minimal Risk Level 

MS mass spectrometry 

NAAQS National Ambient Air Quality Standard 

NAS National Academy of Science 

NATICH National Air Toxics Information Clearinghouse 

NATO North Atlantic Treaty Organization 

NCE normochromatic erythrocytes 

NCEH National Center for Environmental Health 

NCI National Cancer Institute 

ND not detected 

NFPA National Fire Protection Association 

ng nanogram 

NHANES National Health and Nutrition Examination Survey 

NIEHS National Institute of Environmental Health Sciences 

NIOSH National Institute for Occupational Safety and Health 

NIOSHTIC NIOSH's Computerized Information Retrieval System 

NLM National Library of Medicine 

nm nanometer 

nmol nanomole 

NOAEL no-observed-adverse-effect level 

NOES National Occupational Exposure Survey 

NOHS National Occupational Hazard Survey 

NPD nitrogen phosphorus detection 

NPDES National Pollutant Discharge Elimination System 

NPL National Priorities List 

NR not reported 

NRC National Research Council 

NS not specified 

NSPS New Source Performance Standards 

NTIS National Technical Information Service 

NTP National Toxicology Program 

ODW Office of Drinking Water, EPA 

OERR Office of Emergency and Remedial Response, EPA 

OHM/TADS Oil and Hazardous Materials/Technical Assistance Data System 

OPP Office of Pesticide Programs, EPA 

OPPT Office of Pollution Prevention and Toxics, EPA 

OPPTS Office of Prevention, Pesticides and Toxic Substances, EPA 

OR odds ratio 

OSHA Occupational Safety and Health Administration 

OSW Office of Solid Waste, EPA 

OTS Office of Toxic Substances 

OW Office of Water 
PBBs and PBDEs C-4 

APPENDIX C 

OWRS Office of Water Regulations and Standards, EPA 

PAH polycyclic aromatic hydrocarbon 

PBPD physiologically based pharmacodynamic 

PBPK physiologically based pharmacokinetic 

PCE polychromatic erythrocytes 

PEL permissible exposure limit 

pg picogram 

PHS Public Health Service 

PID photo ionization detector 

pmol picomole 

PMR proportionate mortality ratio 

ppb parts per billion 

ppm parts per million 

ppt parts per trillion 

PSNS pretreatment standards for new sources 

RBC red blood cell 

REL recommended exposure level/limit 

RfC reference concentration 

RfD reference dose 

RNA ribonucleic acid 

RQ reportable quantity 

RTECS Registry of Toxic Effects of Chemical Substances 

SARA Superfund Amendments and Reauthorization Act 

SCE sister chromatid exchange 

SGOT serum glutamic oxaloacetic transaminase 

SGPT serum glutamic pyruvic transaminase 

SIC standard industrial classification 

SIM selected ion monitoring 

SMCL secondary maximum contaminant level 

SMR standardized mortality ratio 

SNARL suggested no adverse response level 

SPEGL Short-Term Public Emergency Guidance Level 

STEL short term exposure limit 

STORET Storage and Retrieval 

TD50 toxic dose, 50% specific toxic effect 

TLV threshold limit value 

TOC total organic carbon 

TPQ threshold planning quantity 

TRI Toxics Release Inventory 

TSCA Toxic Substances Control Act 

TWA time-weighted average 

UF uncertainty factor 

U.S. United States 

USDA United States Department of Agriculture 

USGS United States Geological Survey 

VOC volatile organic compound 

WBC white blood cell 

WHO World Health Organization 
PBBs and PBDEs C-5 

APPENDIX C 

> greater than 

≥ greater than or equal to 

= equal to 

< less than 

≤ less than or equal to 

% percent 

α alpha 

β beta 

γ gamma 

δ delta 

μm micrometer 

μg microgram 

q1

*

 cancer slope factor 

– negative 

+ positive 

(+) weakly positive result 

(–) weakly negative result 

PBBs and PBDEs D-1 

APPENDIX D. INDEX 

absorbed dose ....................................................................................................................................245, 253, 297, 303 

Absorption .........................................................................................................................................242, 243, 314, 335 

ACGIH ..............................................................................................................................................................480, 485 

Adipose tissue............................................................................................................................................224, 423, 471 

adrenal gland .............................................................................................................................................................185 

adrenals.................................................................................................................. 48, 75, 185, 210, 248, 250, 251, 253 

adsorbed............................................................................................. 379, 382, 395, 396, 397, 398, 400, 401, 404, 406 

adsorption .................................................................................................................. 357, 395, 397, 398, 399, 400, 460 

aerobic ....................................................................................................................... 403, 406, 407, 408, 460, 477, 478 

Agency for Toxic Substances and Disease Registry................................................ 8, 9, 19, 20, 45, 247, 270, 288, 317 

Ah receptor ............................................................................................................................................................25, 33 

AHH ..................................................................................................................................................271, 272, 274, 278 

Air.............................................................................................. 357, 362, 387, 401, 409, 410, 411, 453, 475, 480, 483 

ambient air.........................................................................................................................381, 409, 410, 450, 460, 461 

anaerobic.................................................................................................................... 381, 403, 406, 407, 408, 460, 478 

anemia ...............................................................................................................................................................171, 322 

antiestrogenic.....................................................................................................................................282, 286, 288, 301 

Arctic ......................................................................................... 397, 410, 419, 430, 431, 432, 438, 439, 440, 453, 456 

ATSDR.................................................................................. 8, 9, 19, 20, 35, 36, 60, 61, 308, 316, 317, 457, 476, 479 

Bioaccumulation................................................................................................................................................400, 461 

Bioavailability ...........................................................................................................................................................460 

bioconcentration factor ......................................................................................................................................396, 399 

Biodegradation ..................................................................................................................................................382, 407 

Biomagnification .......................................................................................................................................................400 

biomarkers ......................................................................................................... 293, 303, 307, 308, 333, 334, 467, 477 

Birds ..........................................................................................................................................................................441 

blood cell .......................................................................................................................28, 43, 171, 172, 196, 300, 330 

blood cells......................................................................................................................................28, 43, 196, 300, 330 

Breast milk.........................................................................................................................................334, 419, 422, 449 

BZ numbers .......................................................................................................................................................346, 354 

Cancer.................................................................... 6, 16, 32, 44, 80, 186, 216, 217, 222, 226, 236, 324, 353, 356, 485 

Carcinogenic..............................................................................................................................................................210 

Carcinogenicity............................................................................................................................................45, 480, 483 

carcinoma .................................................................................................... 44, 210, 219, 220, 221, 222, 225, 282, 325 

carcinomas.........................................................................................................................................32, 40, 45, 53, 219 

cardiovascular..............................................................................................................................70, 167, 168, 169, 229 

Cardiovascular Effects.........................................................................................................................................70, 167 

Children ..................................................6, 7, 16, 17, 208, 291, 292, 294, 302, 313, 339, 340, 422, 454, 455, 464, 465 

chromosomal aberrations...................................................................................................................................241, 326 

clearance.............................................................................. 41, 176, 181, 269, 279, 281, 284, 291, 299, 304, 309, 338 

computer monitors...........................................................................................................................................1, 14, 375 

computer-disassembly ...............................................................................................................................................447 

congener ........................14, 24, 33, 40, 43, 47, 57, 59, 60, 75, 185, 190, 198, 202, 204, 208, 215, 216, 231, 243, 245, 

 248, 249, 251, 252, 255, 256, 257, 260, 263, 269, 272, 274, 275, 276, 278, 279, 282, 284, 289, 

 291, 300, 306, 307, 330, 332, 334, 337, 338, 346, 354, 358, 361, 366, 397, 402, 405, 407, 413, 

 415, 424, 427, 430, 432, 433, 436, 437, 441, 446, 447, 448, 449, 452, 453, 454, 455, 459, 460, 

 462, 463, 466, 467, 469, 477 

death ..................................7, 28, 30, 31, 60, 62, 80, 163, 164, 165, 166, 167, 168, 170, 172, 174, 177, 182, 192, 193, 

 226, 282, 310, 316, 322 

Debromination...........................................................................................................................................................407 

DecaBDE.........................14, 16, 37, 38, 40, 46, 59, 173, 179, 187, 194, 258, 259, 264, 278, 280, 312, 314, 366, 369, 

 386, 389, 399, 404, 406, 410, 413, 432, 433, 442, 447, 449, 455, 461 
PBBs and PBDEs D-2 

APPENDIX D 

deoxyribonucleic acid................................................................................................................................237, 238, 240 

Dermal Effects.............................................................................................................................................76, 190, 231 

Distribution................................................................................................................ 246, 249, 251, 292, 335, 336, 371 

DNA .......................................................................................................... 198, 237, 238, 239, 240, 241, 275, 303, 326 

dopamine ...........................................................................................................................................201, 282, 311, 316 

Drinking water...........................................................................................................................................................483 

effluents .............................................................................................................................376, 392, 412, 425, 426, 430 

electron capture detection..................................................................................................................471, 472, 474, 475 

electron-capture negative ionization..........................................................................................................................468 

electronic enclosures............................................................................................................................................12, 374 

Elimination ........................................................................................................................................248, 254, 261, 269 

elimination half-time .................................................................................................................................................254 

elimination half-times................................................................................................................................................254 

elimination rate..........................................................................................................................................................304 

endocrine ....................................................................... 31, 75, 183, 231, 275, 285, 286, 287, 288, 301, 322, 341, 343 

Environmental Protection Agency...........................................................................................1, 11, 342, 353, 356, 485 

EPA ...........................................1, 6, 8, 11, 16, 19, 33, 45, 61, 188, 216, 285, 302, 316, 317, 341, 342, 353, 356, 363, 

 365, 367, 370, 376, 379, 383, 386, 388, 389, 393, 394, 395, 452, 456, 457, 458, 466, 467, 476, 

 480, 481, 482, 483, 485, 486 

erythema ..................................................................................................................................29, 30, 77, 191, 232, 233 

estrogen receptor..................................................................................................................34, 217, 282, 286, 288, 301 

FDA...........................................................................................................................................8, 19, 58, 417, 480, 485 

fetal tissue................................................................................................................................................7, 17, 250, 254 

fetus ............................................................................................................. 7, 16, 17, 55, 285, 291, 292, 297, 312, 333 

FireMaster..........................2, 24, 26, 27, 28, 29, 30, 31, 32, 33, 34, 57, 58, 77, 80, 163, 164, 166, 168, 169, 170, 171, 

 172, 173, 175, 176, 177, 180, 181, 182, 183, 184, 185, 191, 192, 193, 196, 197, 198, 200, 201, 206, 

 209, 210, 211, 219, 220, 222, 223, 229, 232, 235, 236, 237, 238, 240, 248, 249, 254, 255, 256, 257, 

 269, 272, 273, 274, 275, 286, 287, 294, 295, 310, 321, 322, 325, 328, 336, 337, 346, 353, 355, 358, 

 359, 367, 374, 379, 383, 387, 394, 396, 402, 403, 408, 420, 421, 457, 474 

flame-ionization detector...........................................................................................................................................468 

flexible polyurethane foam (FPUF)...................................................................................................................372, 384 

Food and Drug Administration..........................................................................................................8, 19, 58, 216, 485 

FSH....................................................................................................................................................................286, 295 

gas chromatography........................................................................................... 259, 404, 467, 468, 471, 472, 474, 475 

Gastrointestinal Effects........................................................................................................................................71, 169 

gel-permeation chromatography................................................................................................................................470 

general population ...........................3, 8, 14, 18, 24, 25, 37, 39, 59, 292, 302, 303, 306, 312, 313, 315, 334, 381, 382, 

 420, 450, 451, 452, 453, 456, 459, 462, 463, 464 

genotoxic .......................................................................................................................................40, 60, 237, 241, 326 

Genotoxicity ..............................................................................................................................................238, 239, 326 

Great Lakes..........................47, 48, 49, 50, 59, 62, 69, 70, 71, 72, 73, 74, 75, 76, 77, 78, 79, 242, 246, 321, 324, 340, 

 341, 354, 366, 368, 373, 374, 398, 410, 412, 417, 418, 420, 427, 441, 458, 486 

groundwater......................................................................... 13, 381, 382, 385, 392, 393, 397, 398, 400, 460, 461, 480 

growth retardation........................................................................................................................................32, 209, 328 

half-life .................................................39, 248, 249, 251, 262, 269, 287, 296, 303, 304, 305, 312, 337, 396, 401, 404 

Hematological Effects .........................................................................................................................................71, 171 

Hepatic Effects ................................................................................................................................26, 43, 72, 174, 229 

Hepatocellular carcinomas.........................................................................................................................................219 

high impact polystyrene (HIPS) ................................................................................................................373, 374, 376 

homolog.....................................................................................................................................247, 345, 354, 361, 469 

humic acid .........................................................................................................................................................404, 468 

hydroxyl radical.........................................................................................................................................................401 

IARC ........................................................................... 6, 16, 33, 45, 353, 355, 357, 367, 379, 383, 458, 480, 485, 486 

immune system................................................................................................................5, 16, 196, 197, 300, 329, 341 

immunological..........25, 28, 40, 44, 60, 77, 78, 194, 195, 196, 197, 198, 199, 235, 274, 295, 317, 329, 330, 333, 339 

incineration................................................................................................ 2, 3, 375, 376, 385, 387, 388, 389, 394, 458 

indoor air ...................................................................................................................................................................452 
PBBs and PBDEs D-3 

APPENDIX D 

Integrated Risk Information System..........................................................................................................................485 

IRIS ...........................................................................................................................................................480, 485, 486 

IUPAC...............................................................................................................................................346, 347, 354, 359 

Kow ...........................................................................................................................................357, 361, 362, 399, 400 

landfills.............................................................................. 375, 376, 382, 384, 385, 388, 393, 394, 432, 458, 459, 481 

LD50..........................................................................................................................................163, 164, 165, 166, 310 

leukemia ..............................................................................................................................33, 163, 218, 220, 221, 225 

liquid chromatography...............................................................................................................404, 473, 474, 475, 476 

lymphoreticular......................................................................................................................................................77, 78 

mass spectrometry .............................................................................................................................467, 472, 474, 475 

melanoma ..................................................................................................................................................................224 

menstrual ...............................................................................................................................31, 35, 206, 287, 327, 479 

Metabolism................................................................................................................................................255, 335, 337 

Michigan Chemical Corporation ............................................... 255, 346, 367, 374, 387, 392, 394, 396, 412, 413, 417 

micronuclei........................................................................................................................................................237, 238 

milk............................3, 4, 6, 7, 13, 15, 16, 17, 18, 24, 28, 37, 39, 46, 58, 59, 209, 211, 242, 247, 255, 263, 269, 284, 

 292, 296, 297, 302, 304, 306, 312, 333, 334, 336, 375, 383, 416, 419, 421, 440, 442, 445, 448, 449, 

 454, 455, 463, 464, 470, 472, 473, 477 

MRL ............................26, 27, 31, 32, 34, 35, 47, 49, 50, 51, 52, 53, 54, 55, 56, 61, 76, 178, 179, 183, 188, 216, 320, 

 321, 322, 323, 324, 325, 327, 328, 479 

Musculoskeletal Effects.............................................................................................................................................173 

National Human Adipose Tissue Survey...................................................................................................................447 

National Priorities List.............................................................................................................................1, 11, 379, 383 

National Toxicology Program .................................................................................................5, 33, 224, 317, 466, 485 

neonatal ................................................................... 41, 42, 56, 202, 204, 207, 241, 252, 281, 297, 298, 326, 331, 340 

neoplasm..............................................................................................................................................................32, 221 

neoplastic............................................................................................. 32, 40, 44, 45, 53, 178, 219, 223, 224, 225, 325 

neurobehavioral ..............................39, 42, 188, 199, 202, 205, 212, 285, 297, 299, 316, 325, 330, 331, 332, 340, 343 

neurodevelopmental......................................................................... 25, 40, 42, 201, 215, 297, 298, 299, 313, 331, 333 

neurotransmitter.........................................................................................................................................................282 

nicotinic receptor .......................................................................................................................................................282 

nicotinic receptors......................................................................................................................................................282 

non-Hodgkin’s lymphoma...........................................................................................................................44, 224, 333 

norepinephrine...........................................................................................................................................................201 

NTP .......................................5, 33, 35, 36, 44, 45, 47, 52, 53, 163, 164, 165, 166, 167, 168, 169, 170, 171, 172, 173, 

 174, 176, 177, 178, 181, 182, 183, 184, 185, 186, 191, 192, 193, 194, 196, 197, 199, 200, 201, 

 206, 207, 210, 219, 220, 221, 222, 224, 225, 226, 237, 239, 240, 241, 244, 250, 258, 264, 278, 

 286, 290, 295, 296, 299, 300, 301, 307, 309, 317, 321, 322, 324, 325, 326, 327, 329, 330, 331, 

 334, 335, 343, 457, 466, 476, 483, 485, 486 

nuclear .......................................................................................................................................................................275 

Occupational Safety and Health Administration .............................................................................................8, 19, 485 

OctaBDE.................................................................................................... 172, 178, 187, 194, 233, 366, 369, 373, 388 

ocular effects ...................................................................................................................25, 76, 77, 192, 233, 234, 320 

odds ratio ...................................................................................................................................................................217 

OSHA ......................................................................................................................................................8, 19, 480, 485 

Partition coefficient ...........................................................................................................................................357, 362 

partition coefficients..................................................................................................................................................399 

PBPD.........................................................................................................................................................................266 

PBPK................................................................................................................. 266, 267, 268, 269, 302, 337, 341, 342 

PentaBDE .................................................................................................................. 179, 180, 187, 366, 368, 369, 406 

pharmacodynamic......................................................................................................................................................266 

Pharmacokinetic ........................................................................................................................................248, 266, 269 

pharmacokinetics.......................................................................................................................................292, 293, 342 

Photolysis ..................................................................................................................................................401, 408, 459 

placenta.................................................................................. 4, 7, 15, 17, 247, 249, 292, 302, 306, 333, 336, 421, 463 

placental tissue...........................................................................................................................................................455 

polybrominated dibenzofurans ..................................................................................................................358, 365, 375 
PBBs and PBDEs D-4 

APPENDIX D 

polybrominated-p-dibenzodioxins.............................................................................................................................365 

polyvinyl chloride (PVC) ..........................................................................................................................................373 

publicly owned treatment works (POTWs) .......................................................................................................375, 458 

Pyrolysis ............................................................................................................................................................358, 387 

rate constant.......................................................................................................................................................269, 304 

Regulations......................................................................................................................................8, 19, 480, 483, 485 

Renal Effects .......................................................................................................................................................73, 180 

retention..................................................................................... 33, 39, 46, 60, 201, 251, 252, 258, 264, 299, 309, 470 

RfD ............................................................................................................................................................480, 485, 486 

rigid polyurethane......................................................................................................................................................373 

sediments ............................................................................. 13, 341, 392, 397, 399, 403, 408, 413, 415, 416, 432, 462 

sewage sludges ..........................................................................................................................................................375 

SGOT...................................................................................................................................................26, 174, 177, 310 

solubility.................................................................................................................... 244, 251, 258, 264, 364, 398, 400 

SRBC.................................................................................................................................................196, 197, 198, 278 

Superfund ..........................................................................................................................................................342, 386 

T3...................................26, 35, 41, 47, 49, 50, 51, 54, 55, 75, 183, 184, 185, 187, 188, 189, 190, 211, 214, 215, 279, 

 280, 281, 283, 284, 286, 288, 289, 290, 291, 295, 298, 300, 301, 307, 309, 312, 323, 332, 334, 338, 340 

T4...............................25, 26, 34, 35, 41, 47, 50, 51, 54, 55, 74, 75, 183, 184, 185, 186, 187, 188, 189, 210, 211, 214, 

 215, 278, 279, 280, 281, 283, 284, 286, 289, 290, 291, 295, 297, 298, 299, 300, 307, 309, 312, 316, 

 323, 332, 334, 335, 338, 340, 479 

televisions......................................................................................................................................................1, 375, 452 

textiles ......................................................................................................... 1, 23, 37, 57, 373, 374, 386, 392, 432, 451 

thyroid ...................................5, 7, 15, 17, 25, 26, 27, 33, 34, 35, 39, 40, 41, 42, 44, 47, 48, 49, 50, 51, 52, 53, 54, 55, 

 56, 74, 75, 183, 184, 185, 186, 187, 188, 189, 190, 206, 210, 211, 214, 215, 222, 225, 231, 

 243, 247, 275, 278, 279, 280, 281, 283, 284, 285, 286, 288, 289, 290, 291, 295, 297, 298, 

 299, 301, 307, 309, 311, 312, 313, 316, 320, 321, 322, 323, 324, 325, 333, 334, 335, 338, 

 339, 340, 341, 479 

thyroid stimulating hormone..........................................................................................................................41, 75, 185 

toxicokinetic ..................................................................................................................39, 57, 299, 309, 313, 337, 338 

toxicokinetics.......................................................................................................................................46, 241, 297, 337 

triiodothyronine ...........................................................................................................26, 183, 279, 283, 288, 289, 301 

TSH ......................................................41, 49, 50, 51, 75, 184, 185, 187, 188, 189, 190, 215, 278, 290, 298, 332, 334 

Tumors.........................................................................................................................................................................33 

ultraviolet...........................................................................................................................................................173, 404 

vapor phase................................................................................................................................382, 396, 398, 402, 453 

vapor pressure.................................................... 329, 357, 361, 363, 364, 384, 387, 389, 396, 397, 398, 453, 457, 463 

volatility.............................................................................................................................389, 395, 397, 398, 400, 470 

volatilization.............................................................................................. 323, 384, 387, 389, 395, 396, 397, 398, 459 

water solubility .......................................................................................... 355, 357, 361, 363, 385, 396, 397, 398, 400 

weanling .......................................................................................... 47, 52, 54, 175, 179, 186, 187, 207, 298, 300, 321 

weanlings...................................................................................................................................................207, 298, 340 

WHO .............................................58, 59, 346, 354, 355, 356, 360, 361, 363, 365, 368, 372, 374, 376, 381, 383, 399, 

 421, 453, 455, 456, 463, 464, 472, 473, 475, 476 

World Trade Center.............................................................................................................................................13, 425