## TOXICOLOGICAL PROFILE FOR POLYBROMINATED BIPHENYLS AND POLYBROMINATED DI PHENYL ETHERS

U.S. DEPARTMENT OF HEALTH AND HUMAN SERVICES Public Health Service Agency for Toxic Substances and Disease Registry

September 2004PBBs 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 30333PBBs and PBDEs vii



### QUICK REFERENCE FOR HE ALTH 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 informati on on a substance. Health care providers treating patients potentially exposed to hazardous substances wi ll 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 subs tance. 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, re productive), 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 sec tion are necessarily observed in the clinical setting. Please refer to the Public Health Statem ent 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 th e 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 or dered 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 desc ribed, 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 th at involve hazardous materials. Volume III— Medical Management Guidelines for Acute Chemical Exposure s—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 assist ance by investigating problems of health and safety in the workplace, recommends standard s 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 e nvironmental agents on human health and well-being. Contact: NIEH S, 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 o ccupational 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 sp ecializing 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 FO LLOWING ATSDR INTERNAL REVIEWS:

1. Health Effects Review. The Health Effect s Review Committee examines the health effects chapter of each profile for consistency and accur acy 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 recommendati ons for derivation of MRLs.

3. Data Needs Review. The Research Implementa tion 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 polybrominat ed 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 Vete rinary Medicine, Oregon State University, 105 Magruder Hall, Corvallis, Oregon;

3. Christopher Metcalfe, Ph.D., Professor, Enviro nmental 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 a nd Risk Assessment Practice Director, Montgomery, Watson, and Harza, Sacramento, California.

These experts collectively have knowledge of polybr ominated biphenyls and polybrominated diphenyl ether's physical and chemical properties, toxicokineti cs, key health end points, mechanisms of action, human and animal exposure, and qua ntification of risk to humans. All reviewers were selected in conformity with the conditions for peer review sp ecified 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 al so 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

............................................................................................................................... ...................37PBBs 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 Pharmacokine tic (PBPK)/Pharmacodynamic (PD) Models................266 5.5 MECHANISMS OF ACTION ..................................................................................................................269

5.5.1 Pharmacokinetic Mechanisms .............................................................................................. ...269 5.5.2 Mechanisms of Toxicity .................................................................................................. ........270PBBs 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 .............................................................................................................................37 0 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 ........................................................................................4 01 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........................................................................................................ ..........465PBBs and PBDEs xvi

9. ANALYTIC AL 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-1PBBs and PBDEs xvii



### LIST OF FIGURES

5-1. Levels of Significant Exposure to Lo wer 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 Physiologica lly Based Pharmacokinetic (PBPK) Model for a Hypothetical Chemical Substance ............................................................................................... 2 68

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...................................................................... 382PBBs and PBDEs xix



### LIST OF TABLES

5-1. Levels of Significant Exposure to Lo wer 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 Fire Master® BP-6 and FireMaster® FF-1.................................... 359

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

6-7. Physical and Chemical Prope rties 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/m 3 ) of Several PBDEs in Air Samples ......................................................... 413

8-3. Concentrations (ng/g) of Several PBDEs in Sediment and Suspended Particulate Samples....... 416PBBs 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 Se veral 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 Se veral PBDEs in Human Adipose Tissue Samples ..... 446

8-13. Concentrations (ng/g Lipid Weight) of Se veral 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 Determini ng PBBs in Environmental Samples...................................... 476

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

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



## 1. PUBLIC HEALTH STATEMENT—PBBs

This public health statement te lls you about polybrominated biphe nyls (PBBs) and the effects of exposure to PBBs.

The Environmental Protection Agency (EPA) identif ies the most serious hazardous waste sites in the nation. These sites are then placed on the Nati onal 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 larg e area, such as an indus trial plant, or from a container, such as a drum or bot tle, it enters the environment. Such a release does not always lead to exposure. You can be exposed to a s ubstance only when you come in contact with it. You may be exposed by breathing, eating, or dri nking the substance, or by skin contact.

If you are exposed to PBBs, many factors will de termine 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 chem icals 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, te levisions, textiles, and plastic foams, to make

them difficult to burn. Because PBBs were mi xed 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 PB Bs was discontinued in the United States inPBBs AND PBDEs 2

1. PUBLIC HEALTH STATEMENT—PBBs



### 1976. Concern regarding PBBs is mainly relate d to exposures resulting from an agriculture contamination episode that occurred in Mi chigan 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 envir onment 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 unde r 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 agricu lture contamination episode that occurred in Michigan in 1973–1974 led to the ces sation 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 WH EN THEY ENTER THE ENVIRONMENT?

In the past, PBBs entered the air, water, and so il 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 mi llions of farm animals and humans living in Michigan at this time. PBBs entered the e nvironment during the dis posal of contaminated animal feed and animal products during the ag riculture contamination episode. PBBs also entered the environment from PBB-containing wast es that manufacturers disposed of in waste sites. Small quantities of PBBs also entere d 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 de creased 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 burd en 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 le vels 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 c ontaminated 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 worker s indicated that those in P BB industries were exposed to higher levels of PBBs than the general popula tion. 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 incinerati on of materials containi ng PBBs. Exposure in workplaces is no longer likely because PBBs ar e no longer manufactured. People who live near hazardous waste sites that contain PBBs may be e xposed primarily by breathing air that contains

PBBs.

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

1.4 HOW CAN PBBs EN TER 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 throug h 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, th ey can partially change into breakdown products called metabolites. Some metabolites and unchanged P BBs 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. Mo re 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. Wit hout 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 re search animals with care and compassion. Scientists must comply with stri ct animal care guidelines because laws today protect the welfare of research animals.

Much of what is known about the health effect s of PBBs in people comes from studies of ingestion in Michigan in the early-to-mid 1970s, where feed fo r farm animals was accidentally

contaminated with a fire retard ant 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 pa in, loss of appetite, jo int pain, fatigue, andPBBs AND PBDEs 5

1. PUBLIC HEALTH STATEMENT—PBBs



### weakness. However, it could not be clearly establishe d that any of the problems were caused by eating the food contaminated with PBBs. PBBs al so did not cause any definite changes in the livers or immune systems of the Michigan re sidents. 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 b ody weight loss, skin disorders, and nervous system effects, and their livers, kidneys, thyroi d glands, and immune systems we re 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 infe rtile. Most of the eff ects in animals occurred after they ate large amo unts of PBBs for short periods or smaller amounts for several weeks or months. Body weight loss and effects on the live rs, 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 st udies tested animals exposed to PBBs by skin contact. These animals had injuries to the liver and skin. On ly 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 can cer in people who ate food contaminated with PBBs. Rats developed cancer in their livers af ter 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 al so developed in rats and mice that ate smaller

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

of a PBB mixture for several months did not devel op 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 an ticipated to be carcinogens. Similarly, thePBBs 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 cl assified the carci nogenicity of PBBs.

We do not know whether the effects found in anim als 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 foun d in the environment. Long-term exposure to these chemicals has a greater poten tial to cause health effects th an 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 po tential health effects in humans fr om exposures during the period from conception to maturity at 18 years of age.

Children are exposed to PBBs in generally th e 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 ob jects in their mouths, by eating without washing their hands, or similar behavior. Some children al so eat dirt on purpose. It is not possible that children could be exposed to PBBs following tr ansport of the chemical on clothing from the parent’s workplace to the home because P BBs are no longer being produced or used.

Some information on health effects of PBBs in ch ildren is available from studies of the Michigan contamination episode. Symptoms of ill health we re not associated with increased exposure to PBBs, and general neurological examinations di d not show any abnormalities. More detailed studies of physical and neuropsychological deve lopment showed no effects that were clearly related to PBBs among Michigan children exposed during the ep isode. Changes in nerve and

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

1. PUBLIC HEALTH STATEMENT—PBBs



### nursing. Animal studies also f ound that exposure to PBBs during pregnancy or lactation caused changes in thyroid hormone levels in the new born animals and, at high doses, increases in prenatal death and structural birth defects.

As indicated above, children can be exposed to P BBs before birth and from breast milk. PBBs are stored in the mother’s body and can be re leased 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 c oncerns 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 fetu ses and children are more susceptible to PBBs than are adults.

1.7 HOW CAN FAMILI ES 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. Y our doctor might need to ask your state health department to investigate.

Because PBBs are no longer produced or used, th e 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 becaus e they could still contain PBBs. Children should also be discouraged from eati ng the dirt, and careful handwashi ng 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 peoplePBBs 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 populati on, they do not predict whether you will be harmed. Blood tests are the eas iest, safest, and probably the best method for detecting recent or past exposures to large amounts of PBBs. Re sults of such tests should be reviewed and carefully interpreted by physic ians 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 info rmation 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 TH E FEDERAL GOVERNMENT MADE TO PROTECT HUMAN HEALTH?

The federal government develops regulations a nd recommendations to pr otect public health. Regulations can be enforced by law. The EPA, the Occupational Safety and Health Administration (OSHA), and the Food and Dr ug Administration (FDA) are some federal agencies that develop regulati ons 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 fe deral organizations that deve lop 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 valu e that is usually based on levels that affect animals; they are then adju sted to levels that will help protect humans.

Sometimes these not-to-exceed levels differ amo ng 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 regulati ons 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 GE T MORE INFORMATION?

If you have any more questions or concerns abou t 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 occupati onal and environmental health clinics. These clinics specialize in recognizing, evaluating, and tr eating illnesses that resu lt from exposure to

hazardous substances.

Toxicological profiles are al so available on-line at www.atsdr.cdc.gov/toxpro2.html and on CD-ROM. You may request a c opy of the ATSDR ToxProfiles TM 

### CD-ROM by calling the tollfree information and technical assistance number at 1-888-42ATS DR (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: 1-770-488-4178PBBs AND PBDEs 10

1. PUBLIC HEALTH STATEMENT—PBBs



### Organizations for-profit may reque st copies of final Toxicologi cal 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 te lls you about polybrominated diphenyl ethers (PBDEs) and the effects of exposure to PBDEs.

The Environmental Protection Agency (EPA) identif ies the most serious hazardous waste sites in the nation. These sites are then placed on the Nati onal Priorities List (NPL ) and are targeted for long-term federal clean-up activit ies. PBDEs have not been f ound in the 1,647 current or former NPL sites. Although the total num ber of NPL sites evaluated for these substances is not known, the possibility exists that PBDEs may be found in the future as mo re sites are evaluated. This information is important because these sites ma y be sources of exposure and exposure to these substances may harm you.

When a substance is released either from a larg e area, such as an indus trial plant, or from a container, such as a drum or bot tle, it enters the environment. Such a release does not always lead to exposure. You can be exposed to a s ubstance only when you come in contact with it. You may be exposed by breathing, eating, or dri nking 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 expos ed 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 adde d 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 Isra el). Decabromodiphenyl ether (decaBDE) makes up 82% of these products manufactured globally. It s main use is for electronic enclosures, such as television cabinets. Octabromodiphenyl et her (octaBDE) product is used in plastics for business equipment. Pentabromodiphenyl ethe r (pentaBDE) product is used in foam for cushioning in upholstery. PBDEs ha ve not been associated with ac tual health-related effects. Concerns have increased, however, because some of these chemicals (particularly PBDEs in the pentaBDEs) have been found in the environmen t at varying concentrations. Environmental concentrations of lower brominated PBDEs, whic h 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 chemi cals with no known natu ral sources in the environment, except for a few marine organism s that produce forms of PBDEs that contain higher levels of oxygen. Commerci al 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 containi ng a variety of individual brominated diphenyl ether (for PBDEs) components, kno wn 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 wide ly 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 propertie s and uses of PBDEs can be found in Chapters 6 and 7.

2.2 WHAT HAPPENS TO PBDEs WH EN THEY ENTER THE ENVIRONMENT?

PBDEs enter air, water, and so il during their manufacture and us e in consumer products. When PBDEs are suspended in air, they can be present as particles. They eventu ally return to land or

water as the dust settles and are wa shed 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, highPBBs 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 eventu ally 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 de caBDE, are not found in fish at measurable concentrations. In general, the breakdown of PBDE s in soil is very slow, so they may remain in soil for several years. PBDEs bind strongly to soil particles. Rainwa ter is not expected to spread

them much below the soil surf ace; 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 produ cer of the pentaBDE mixture for commerce. However, production pentaBDE and octaBDE for co mmerce soon will begin to be phased out in

the United States and elsewhere. Lower bromin ated PBDEs, such as tetraBDE and pentaBDE congeners, are found throughout the environment and ar e found at low levels in air, sediments, animals, and food. The concentrations of lo wer brominated PBDEs in blood, breast milk, and body fat indicate that most people ar e 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 br ominated 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 pl aces 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 Unite d States have been c onducted to identify the sources of exposure, people appear to be expos ed to the lower brominated congeners of PBDEs by eating food that contains these PBDEs. In the United States, the c oncentration of PBDEsPBBs 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/m 3 

### ), which indicates low levels of exposure of the general population to these PBDE s. 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 a nd 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 ar e 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-contamin ated 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 bromin ated PBDEs have been detected in air samples, indicating that people ca n also be exposed by inhalation. Th e ways that PBDEs might enter and leave your body depend on the chem ical structures of the cong ener components. The higher brominated PBDEs, particularly decaBDE (the ma jor 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 containi ng PBDEs, which could happen at a hazardous waste site, it is highly unlikely th at either lower or higher brom inated PBDEs would pass through your skin into the bloodstream. Once PBDEs ar e in your body, the congeners might partially change into breakdown products called me tabolites. DecaBDE might leave your body unchanged or as metabolites, mainly in the feces a nd in very small amounts in the urine, within a few days. Lower brominated PBDEs, generally tetra-, penta-, and he xaBDE congeners, might stay in your body for many years, stored mainly in body fat. The lower brominated PBDEs alsoPBBs and PBDEs 15

2. PUBLIC HEALTH STATEMENT—PBDEs



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

PBDEs also can enter the bodies of unborn babi es through the placenta. More information on how PBDEs can enter and leave you r 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 chemi cals, animal testing may be necessary. Animal testing may also help identify heal th effects such as cancer or bi rth defects. Without laboratory animals, scientists would lose a basic method for getting information needed to make wise decisions that protect public hea lth. Scientists have the responsi bility to treat research animals with care and compassion. Scient ists must comply with strict animal care guidelines because laws today protect the welfare of research animals.

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

available information is from studies of laborat ory animals. Animal studies indicate that

commercial decaBDE mixtures are generally much less toxic than the products containing lower brominated PBDEs. Because of its very diffe rent toxicity, decaBDE is expected to have relatively little effect on the hea lth of humans. Rats and mice th at 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 w eeks or months developed effects in the liver and in the thyroid. It is specula ted that many of the thyroid effect s of PBDEs are specific to the species of test animals, suggesti ng that they are less likely to occu r 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 ch anges 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-te rm animal studies suggest that some PBDEs might impair thePBBs 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 larg e amounts of decaBDE throughout their lifetime. On the basis of evidence for cancer in animals, decaBDE is clas sified 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 In ternational Agency for Research on Cancer (IARC) have classified the carcinogenicity of any PBDEs.

We don’t know whether the effects found in anim als 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 eff ects 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 octa BDE products are much more likely to cause

health effects than is decaBDE. More inform ation on how PBDEs can affect your health can be found in Chapter 5.

2.6 HOW CAN PBDEs AFFECT CHILDREN?

This section discusses po tential health effects in humans fr om exposures during the period from conception to maturity at 18 years of age.

Children are exposed to PBDEs in generally th e same way as are adults, probably mainly by eating contaminated food. Becau se of their smaller weight, ch ildren’s intake of PBDEs per kilogram (or pound) of body weight ma y be greater than that of adu lts. 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 ex posed. DecaBDE is poorly absorbed in the body and, therefore, is unlikely to be found in breast milk or the fe tus to any significant extent.

Children who live near hazardous waste sites might accidentally eat some PBDEs by puttingPBBs 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 purpos e. It is also possibl e that children could be exposed to PBDEs following transport of the chem ical on clothing from th e parent’s workplace

to the home.

As indicated above, children can be exposed to PBDEs—mainly the lower brominated congeners—both before birth a nd 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 breas t milk. In most cases, however, the benefits of breast-feeding outweigh any risks from exposur e 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 mo re significant if expos ure occurs during the periods before and soon after birt h. That suggests that fetuses a nd 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 FAMILI ES 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. Y our doctor might need to ask your state health department to investigate.

PBDEs are currently in widespread production a nd 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 locat ions. Children who live near ha zardous waste sites should bePBBs and PBDEs 18

2. PUBLIC HEALTH STATEMENT—PBDEs



### discouraged from playing in the dirt near th ese sites because they could contain PBDEs. Children should also be discouraged from eati ng dirt, and careful handw ashing practices should

be followed.

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

commercial PBDE mixtures and of PBDE-contai ning 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 occupationa l health and safety officer at work can tell you whether the products you work with may contain PBDEs and whet her those are likely to be carried home. If this is the case, you should show er and change clothing before l eaving work. Your work clothes should be kept separate from othe r 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 routin e clinical tests, but c ould be ordered by a doctor to detect PBDEs in people exposed to them in the envi ronment and at work. If your PB DE 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 te sts cannot distinguish between recent or past exposures to PBDEs b ecause 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 wh ether you will be harmed. Blood tests are the easiest, safest, and probably the best method fo r detecting recent exposures to large amounts of

PBDEs. Results of such tests should be review ed and carefully interpreted by physicians with a background in environmental and occupational medi cine. Nearly everyone has been exposed to pentaBDE commercial mixtures because they are found throughout the environment. That means people are more likely to have detectab le amounts of the lower brominated PBDEs in their blood, fat, and breast milk. Recent studies have shown that levels of lower brominatedPBBs 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 indi viduals living in Europe.

More information on tests used to determine wh ether 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 TH E FEDERAL GOVERNMENT MADE TO PROTECT HUMAN HEALTH?

The federal government develops regulations a nd recommendations to protect public health. Regulations can be enforced by law. The EPA, the Occupational Safety and Health Administration (OSHA), and the Food and Dr ug Administration (FDA) are some federal agencies that develop regulati ons 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 fe deral organizations that deve lop 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 valu e 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 di ffer among federal organizations because they used different exposure times (an 8-hour workday or a 24-hour day), differe nt 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 regulati ons and recommendations for PBDEs include the following:

EPA requires that companies that tr ansport, store, or dispose of p -bromodiphenyl ether (a particular PBDE compound not found in any co mmercial PBDE product) follow the rules andPBBs and PBDEs 20

2. PUBLIC HEALTH STATEMENT—PBDEs



### regulations of the federal haza rdous 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 regula tions and guidelines for PBDEs, see Chapter 10.

2.10 WHERE CAN I GE T 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 occupati onal and environmental he alth clinics. These clinics specialize in recognizing, evaluating, and tr eating illnesses that resu lt from exposure to

hazardous substances.

Toxicological profiles are al so available on-line at www.atsdr.cdc.gov/toxpro2.html and on CD-ROM. You may request a c opy of the ATSDR ToxProfiles TM 

### CD-ROM by calling the tollfree information and technical assistance number at 1-888-42ATS DR (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: 1-770-488-4178PBBs and PBDEs 21

2. PUBLIC HEALTH STATEMENT—PBDEs



### Organizations for-profit may reque st copies of final Toxicologi cal 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 or ganic 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. Comme rcial production of PBBs began in approximately

1970 and manufacture was discontinued in the United St ates 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 ex ist in 3, 12, 24, 42, 46, 42, 24, 12, 3, and 1 isomers, respectively. The general chemi cal structures of PBBs are similar when viewed in one dimension, as shown below, where m+n = 1–10:



### 1' 2' 3'

4'

5' 6' 1 2 3

4

5 6 Br n Br m

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 molecu les’ 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 a nd 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 PB Bs 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 ag ent. The hexabromobiphenyl mixtures contained varying proportions of di- through octabrominated ho mologues. 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 mixt ures contained a large proportion (47.4–60.0%) of nonabromobiphenyl congeners, whereas commercial decab romobiphenyls contain predominately (96.8%) decabromobiphenyl congener.

Limited data are available on health effects of commercial decabromobiphenyl and octabromobiphenyl mixtures, although the hexabromobiphenyl mixtures Fi reMaster 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 su pplement and mixed with feed that was distributed within the state for several months before being disco vered. Health problems in dairy cattle, reported in the fall of 1973, were the first signs that this epis ode occurred, but the accidental addition of PBBs to

animal feed was not identified as the cau se 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 ingest ed the original PBB mixtures or metabolic products of the PBBs. Based on limited information in dairy cattle and add itional data in laboratory animals as discussed in Section 5.4.2.2, it is reasonable to assume that ma inly 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 environmentall y distributed in Michigan via waste disposal on farms. The general population outside of Michigan could possibly have been exposed to PBBs by thePBBs 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 in to the feed/food chain, and some

information is available on health effects in PPB-e xposed chemical workers. Although the human studies consist largely of observations on groups that are not well defined and lack accura te 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. T hus 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, gast rointestinal, hepatic, derm al, 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 hum ans include neurodevelopmental effects in exposed children and immunological changes. Physical exam inations and laboratory tests have shown few abnormalities that corresponded to the complaints, and pr evalences of symptoms have not been correlated with serum PBB levels. However, the possibilit y 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, neur ological, and developmental. Other effects of oral exposure to PBBs include decr eased 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 si gnificant 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 alte red vitamin A homeostasis, thymic atrophy, dermal and ocular effects (e.g., chlo racne 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. Eff ects in workers exposed to unspecified PBBs and/or decabromobiphenyl included increased serum thyrotr opin, low or borderline low serum thyroxine (T 4 ),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 decreas es in serum levels of serum T 4 and serum triiodothyronine (T 3 ) to histological and ultrastructural cha nges in the follicles. The preponderance

of these studies tested FireMaster FF-1 or FireMast er 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-dur ation studies at doses as low as 3 mg/kg/day (reduced

serum levels of T 4 hormone), but not at 1 mg/kg/day, and in ra ts 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 chr onic 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 implicat ions are not as clear. The intermediate-duration LOAELs for thyroid and hepatic effects are also comp arable to each other, but neither of these LOAELs is suitable for an intermediate MRL because reproduc tive 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 documen ted 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 notdemonstrate 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 de rmal routes. Although the available studies on liver effects in humans are largely inconclusive, the anim al 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 induc tion and liver enlargement to fatty changes and

necrosis have been observed in rodents and other labo ratory animal species exposed orally to FireMaster PBBs in acute-, intermediate-, and/or chronic-dura tion 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 si milar 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 we re reported in the liver of rats exposed to an octabromobiphenyl mixture in acute- a nd 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 mixtur e exposure. Acute dermal exposure to commercial mixtures of hexabromobiphenyl, but not octabromobiphe nyl, has been reported to produce gross necrotic changes in the liver of rabbits. Hepatocyte enlarg ement and degenerative changes (vacuoles or necrosis) are the most sensitive adverse hepatic effects that ha ve been observed in the acute-, intermediate-, and

chronic-duration oral studies with FireMaster PB Bs. 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 expo sure 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 LOAE L is a suitable basis for a chronic MRL, however,

due to decreased survival at the sa me 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 epit helial cells, and some PBB-induced epithelial lesions resemble those produced by vitamin A deficiency. B ecause 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 Michig an contamination episode have been reported by some investigators. Others have not been able to confirm these findings. However, it is clear that noPBBs and PBDEs 28

3. RELEVANCE TO PUBLIC HEALTH—PBBs

correlation can be established between altered immune parameters and P BB levels in serum. Some have suggested that PBBs associated with white bl ood 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. C ontinuous 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 th ymus 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 appare nt, however, that some of these effects are only seen at PBB levels that cause overt toxicity. Steroids are known to influence the immune response. Cort icosterone levels were elevated in plasma of mice that were exposed to Fi reMaster 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 a pproached death at this dose. Based on the data available, it is difficult to suggest any particular sp ecies as the most sensitive. This is because different studies usually examined different end points, using di fferent 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 lim ited data on humans are largely inconclusive, PBBs have altered immune responses in a variety of anim al 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 li mited number of animal studies both suggest that exposure to PBBs may cause subtle effects on ne uropsychological performance and development in humans. Symptoms of neurological effects, includi ng fatigue, weakness, and decrements in the capacity to perform physical or intellectual work, were re ported 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 betweenPBBs 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 P BBs via inhalation and/or dermal contact.

Examinations of a small number of childre n (19) believed to have been exposed in utero or in early infancy during the peak of the Mich igan 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 activ ity 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 offspri ng of female rats and fema le 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 locomoti on, cliff avoidance, cage emergence, and open-field activity were found in offspring of rats that were ex posed 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 prevalen ces 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 Mi chigan 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 studi es 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 subcutane ous 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 thePBBs 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 atroph y 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 co nsistent and indicate that although PBBs can cause local responses such as irritation by direct dermal co ntact, 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/m 3 ) 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. Hist opathological 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 pr eceding death. Effects on body weight have been observed in single dose, intermediate- and chronic-dur ation oral studies with rats, mice, guinea pigs, mink, monkeys, and/or cows. Changes in body weig ht were also observed in rabbits following acute dermal exposure to commercial mixtures hexabromobi phenyl, but not octabromobiphenyl, suggesting that the syndrome is independent of exposure route a nd is a potential effect of PBBs in humans.

Reproductive Effects. A limited amount of data is availabl e regarding the reproductive effects of PBBs in humans. The distribution of sperm counts, sperm motility, and sperm morphology wasinvestigated in a small number (50) of male Michigan residents who ingested food produced on PBBcontaminated farms or who worked in a PBB manufac turing company. The study found no evidence for PBB-related effects compared with a putatively une xposed control group. No relationship was found between serum PBB levels and frequency and duratio n 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 FireMast er 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 dur ing 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 devel opmental toxicity following exposures of <1 year precludes deriving an MRL for intermediate-duration e xposure. Implantation was completely blocked in

40–67% of female rats treate d with gavage dose levels ≥ 28.6 mg/kg on alternate days between gestation days 0 and 14. Alterations in male reproductive orga ns were observed at doses that caused death in male rats (necrosis of the epithelial lining of the ductu s 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 F 1 or F 2 generations of female pa rental 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 ca nnot 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 Mich igan counties with no quarantined farms did not clearly establish or refute the possibility that the Mi chigan 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 duri ng gestation and in monkeys after exposure before conception and during pregnancy. Because th e 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 toPBBs and PBDEs 32

3. RELEVANCE TO PUBLIC HEALTH—PBBs

derive an MRL for chronic-duration exposure. Structur al 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 terat ogenic in rats. Studies with FireMaster FF-1 and FireMaster BP-6 found that body weight gain was redu ced in the offspring of rats and mice after exposure during gestation, in rat offspring after exposure duri ng 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, hepato cyte 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. Pe rformance deficits in tests of operant behavior were observed in offspring of rats and mice after oral e xposure to FireMaster FF-1 or FireMaster BP-6 during pregnancy and lactation.

Although the available human data re garding developmental effects of P BBs are inconclusive, the results from animal studies strongly suggest that PBBs ma y cause mild to severe de velopmental 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 Fi reMaster 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 perinata l 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 shorte r-duration exposures of up to 6 months followed by an observation period of up to 2 years. Based on fi ndings in male rats and mice of both sexes in thisPBBs and PBDEs 33

3. RELEVANCE TO PUBLIC HEALTH—PBBs

study, there is some evidence that combined perina tal and adult dietary exposure to FireMaster FF-1 enhanced the susceptibility of hepatocellular neopl asms 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 wa s inconclusive in mice in both National Toxicology Program bioassays. Equivocal increases in incide nces of mononuclear cell leukemia were observed in adult-only exposed rats in the NTP chronic study, and combined perinatal a nd adult exposure showed no significant increase. Combined anal ysis of the incidences of this leukemia in the adult-only, perinatal only, and combined perinatal a nd 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 administ ration 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 evid ence of carcinogenesis in humans. NTP concluded that PBBs are reasonably anticipat ed 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 wh ich constituents of the PBB mixtures are carcinogenic and the congener profile to which pe ople may be exposed environmentally is likely to be different from the original PBB source, it is assumed that PBB mixt ures 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 rodent s 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 comm on mechanism of toxic action that involves a specific cytosolic molecular receptor (Ah receptor) (see Chapte r 5, Section 5.5.2, Mechanisms of Toxicity). Although numerous factors can influence the toxici ty 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 magn itude of the response that is initiated by its binding with the Ah receptor. The binding affinity, in turn, is determ ined by the substitution pattern of the congener, andPBBs and PBDEs 34

3. RELEVANCE TO PUBLIC HEALTH—PBBs

many of the most toxic congeners resemble the stru ctural 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 intermediate s (e.g., arene oxides) that can form potentially toxic covalently bound substrate-macromolecular adducts.

People are environmentally exposed to PBB mixtur es of different congeneric composition than the original commercial PBB products. Although the toxi city 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 me chanisms, 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 expos ures 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 he alth 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. Insu fficiencies in the human inhalation data include mixedchemical and unquantified exposures. The animal inha lation database is limited by inadequately reported studies and lack of any information on the mixtures lik ely 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 T 4 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 wasPBBs 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 T 4 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 th e designation of the NOAEL and LOAEL values. The data in the manuscript are presented graphically, w ith the animal numbers presented as a range (8– 11 animals/group); thus, an ANOVA could not be perfo rmed from published report. However, using the graphical data, the change in plasma T 4 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 contrai ndicated by the lack of appropriate statistical analysis.

Decreased serum T 4 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 T 3 and T 4 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-durati on 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 lo west dose tested in the database. This dose (0.012 mg/kg/day) caused increased menstrual cycl e duration and implantation bleeding after 6–7 months of exposure and fetal deaths (feta l abortion and stillbirth) after ≈ 1 year of exposure in monkeys, with surviving infants having decreased birth weight and d ecreased 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 de velopmental 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-dur ation exposure. Derivation of an MRL for chronic oral exposure is similarly precluded by the serious developmental e ffect (stillbirth) that occurred following exposuresPBBs 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 L OAELs into less serious and serious effects is

discussed in the introduction to Section 5.2.PBBs and PBDEs 37



### 4. RELEVANCE TO PU BLIC HEALTH—PBDEs

4.1 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO PBDEs IN THE UNITED STATES

Polybrominated diphenyl ethers (PBDEs) are brom inated 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. Therefor e, they have the potential to migrate from the plastic

matrix into the environment when conditions are id eal. Production of PBDEs began in the 1970s and has continued to the present. Concern for the possible hea lth effects of PBDEs has heightened recently due to evidence that components of pentabromodipheny l 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 populatio n is exposed to low levels of primarily lower brominated (e.g., tetra- and penta-) BDEs. For exam ple, the median concentrations of BDE 47, BDE 153, BDE 183, BDE 209, and total PBDEs (sum of four c ongeners) were 0.63, 0.35, 0.17, <1, and 2.2 ng/g lipid weight, respectively. DecaBDE was found at le vels above the limit of quantification (1 pmol/g lipid). Environmental concentrations of lower brom inated PBDEs (i.e., tetraBDE and pentaBDE) appear to be leveling off in Europe, but appear to be incr easing in certain areas in Canada and the United States, although data are too sparse to make broad statemen ts regarding trends. The concentration of total PBDEs in air ranges from 5.5 pg/m 3 in rural environments to 52 pg/m 3 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 so ils 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 et her). Monobrominated structures (i.e., one bromine atom attached to the molecule) are often include d when describing PBDEs. There are 209 different molecular combinations, or congeners, that are possi ble 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 decabro minated), with each homologous group containing one or more isomers. The mono-, di-, tri-, tetra- , penta-, hexa-, hepta-, oc ta-, 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

Br n Br m 

### 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 numb er 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 congeneri c compositions. People are also environmentally exposed to lower brominated PBDEs (e.g., tetra- a nd penta- brominated congeners) due to differential partitioning and transformation of the individual cong eners 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 con tinue 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 l ess 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 ar e controversial. Octa- and pentaBDE commercial mixture components, that have a total of four or less bromines, likely undergo differential partitioningPBBs 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-brominate d tetra- and penta-congeners, pa rticularly 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). Lowe r-brominated tetra- and pent a- congeners have been detected in air samples, including those taken from re mote 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 eff ects 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 toxicokine tic differences appear to be related to the number and location of bromines on the diphenyl oxide molecu le, and correlate with e nvironmental monitoring data indicating that decaBDE has low bioaccumulati on potential. Three-dimensional differences in molecular structure might also contribute to the dissi milar 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 lowe r 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 eff ects mainly in the liver and thyroid, particularlyPBBs 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 e ffects 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 la ter in life. Concern for neurodevelopmental toxicity of PBDEs is further raised by the documented e ffects of lower brominated commercial mixtures on thyroid hormone homeostasis, and critical involvement of thyroid hormones in central nervous system development.. Several acute-duration studies w ith commercial pentaBDE mixtures and the single congener, 2,2’,4,4’-tetraBDE (BDE 47), suggest th at immune suppression might be another important health end point for lower brominated BDEs, alt hough comprehensive immunological evaluations have not been performed on any congener or commercial mixt ure. 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 expo sure-related functional e ffects. Developmental toxicity studies have shown no evidence of teratogeni city of penta- and octaBD Es in rats and rabbits, although fetotoxic effects, including skeletal ossifi cation 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-puri ty 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 effect s 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 wei ght of available genotoxicity evidence indicates that PBDEs are not genotoxic. Dermal application studi es in rabbits showed that decaBDE, octaBDE, and pentaBDE were nonirritating to intact skin and cau sed some erythematous and edematous responses in

abraded skin. DecaBDE was not a skin sensiti zer in humans and octaBDE and pentaBDE were nonsensitizing in guinea pigs, indicating that the PBDE s 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 thyroidPBBs and PBDEs 41

4. RELEVANCE TO PUBLIC HEALTH—PBDEs

stimulating hormone (TSH), low or borderline low serum T 4 , 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 attr ibuted to either chemical. There was no clear association between plasma levels of BDE 47 and thyroid hormone levels (free and total T 3 and T 4 , TSH, free testosterone, follicle-stimulating hormone, luteni zing 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 bromin ated PBDEs and a possible target of toxicity in exposed humans.

Thyroid effects, which mainly included reduced serum T 4 hormone levels and follicular cell hyperplasia, were consistently observed in rats and mice ora lly exposed to lower brominated commercial PBDE mixtures. Accompanying changes in serum TSH levels were not found, and the depression of serum T 4 seems to involve enhanced metabolic formation of h ydroxylated metabolites of PBDEs, which bind with high affinity to thyroid transport proteins. Acute duration studies showed decreases in serum T 4 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 ob served in intermediate-duration studies included thyroid hyperplasia in rats exposed to ≥ 8 mg/kg/day octaBDE for 30 days and reduced serum T 4 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 T 4 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 T 4 or T 3 might vary with species and rats are often regarded as more sensitive than humans. The main basi s for this opinion seems to be studies showing that PBDEs affect binding of thyroid hormones to transthyre tin (TTR), the primary transport protein in rats. Because TTR is not the major transport protein in hu mans, 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 relate d 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 thyr oid 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 T 4 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 misl eading 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 ne urotoxicity 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 oftesting 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, ha bituation, and auditory startle response, although suggestive decreases in fear c onditioning were observed.

Neurobehavioral effects of individua l 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 defi ned critical phase of neonatal brain development, as shown by mild impairments in spontaneous motor beha vior 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 (bod y weight gain, hair growth, day of eyelid and ear opening, day of incisor eruption) and neurobehavior al (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 delaye d sensorimotor development and altered spontaneous

behavior. Based on the limited ava ilable information (most of the studies were reported as abstracts), and considering the known effects of lower brominated PBDEs on thyroid hormone homeostasis and thePBBs and PBDEs 43

4. RELEVANCE TO PUBLIC HEALTH—PBDEs

critical role that thyroid hormones play in developm ent of the central nervous system, neurobehavioral development is a potential effect of concer n 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 ob served hepatic effects includes microsomal enzyme induction, liver enlargement, and degenerative histopathologic altera tions that progress to tumors. Repeated dietary exposure to PB DEs 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 oc taBDE and pentaBDE than decaBDE. For example,

subchronic oral studies in rats showed that comme rcial pentaBDE mixtures were hepatotoxic at doses

≤ 10 mg/kg/day. Increased liver weight and hepatocellula r 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 comme rcial decaBDE caused liver effects only following lifetime exposure to doses that were still very hi gh. 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 ki ller 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, lymphoi d hyperplasia) in rats exposed to ≥ 1,200 mg/kg/day forPBBs and PBDEs 44

4. RELEVANCE TO PUBLIC HEALTH—PBDEs

103 weeks, indicating that it is considerably less im munotoxic than pentaBDE. No studies are available on immunological effects of PBDEs in humans. Du e 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. G estational exposure to a high (1,000 mg/kg/day) but maternally nontoxic dose of decaBDE was fetotoxic in rats as shown by subcut aneous 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 cau sed skeletal ossification variations in rats and rabbits at maternally toxic levels and other indica tions of fetotoxicity at lower doses. Effects of gestational exposure to octaBDE included minima lly 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 animal s, 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/da y, 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, al though the term neoplastic nodule is poorly defined and understood, and is no long er used by NTP to characterize hepatoproliferative lesions in rats. Incidences of hepatocellular ad enoma or carcinoma (combined) were significantly increased in the male mice. Slight ly elevated incidences of thyroid gland follicular cell adenoma orPBBs 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% deca BDE mixture (containing 21.8% nonaBDPO and 0.8% octaDBDPO) for approximately 2 years. No expos ure-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 co mbined 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 carcinogenic ity) 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 cl assified the carcinogenicity of any PBDE mixture.

4.3 MINIMAL RISK LEVELS (MRLs)

PBDEs share some toxicological properties with ot her 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 particul ar, the oxygen bridge of the ether linkage in the diphenyl oxide molecule increases the distance betwee n 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 rotatio n 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 fo r 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 h as 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 e ffects. The assumption that PBDEs share many toxicological characteristics with PCBs also does not consider geometrical differences due to the higherPBBs and PBDEs 46

4. RELEVANCE TO PUBLIC HEALTH—PBDEs

atomic weight and considerably larger molecular vol ume of bromine compared to chlorine (Hardy 2000, 2002b). These differences contribute to dissimilar phy sical/chemical properties that can influence the

relative toxicokinetics and toxicities of the chemic als. 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 fo r 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 mixt ures of different congeneric composition than the original commercial PBDE products. Although the t oxicity 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 di rectly base MRLs for environmental PBDEs. Due to the likelihoods that (1) multiple mechanisms (Ah-receptor-dependent mechanisms, Ah-receptor independent mechanisms, or both) ma y 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 larg ely resistant to environmental degradation, whereas octa- and pentaBDE commercial mixture components are likely to undergo di fferential partitioning and transformation, such that tetra- and pentaBDEs ar e 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 lipi d-rich tissues; decaDBE does not appreciably partition into lipid-rich tissues and has higher rates of metabo lism and elimination. These characteristics seem to

be a function of both the number and location of brom ines 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 si gnificantly less toxic than lower brominated BDE mixtures. For example, subchronic oral studies in rats showed that commercial octa- or pentaBDEPBBs 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 (throm bosis 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 acu te oral exposure to commercial decaBDE had no effect on thyroid hormones, while similar exposur es to commercial octaBDE or pentaBDE mixtures

caused decreases in serum T 4 and T 3 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 dissi milar toxicity and envi ronmental chemistry of decaBDE as discussed above, separate MRLs were derive d 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 brominat ed BDEs is not recommended at this time due to

insufficient information. The inhalation database fo r 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/m 3 for 8 hours/day for 14 consecutive days. The average mass median aerodynamic diameter (MM AD) 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 ocul ar 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 orga ns including thyroid/parathyroid), gross pathology, and histology (21 tissues including nasal turbinates, trachea, lungs, a nd 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 obser ved by the end of each exposure period in rats exposed to ≥ 24 mg/m 3 ; this effect always disappeared by the following morning. Liver weight was significantly increased and hepatic l esions occurred in rats exposed to ≥ 3.7 mg/m 3 . At 3.7 mg/m 3 , the liver lesions consisted of very slight to slight, focal to multifocal cytoplasmic enlargement of the hepatocytes, accompanied by focal acidophilic degenera tion 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 hist ological 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 Un ion 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 fi ve female Crl:CD(SD)IGS BR rats, by nose-only inhalation as a dust aerosol, in me asured concentrations of 0 (air only), 1.0, 10, 110, or 250 mg/m 3 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 repor ted 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 h eart. Exposure-related effects occurred in the liver and nasal cavity. Hepatic effects included increased mean absolute and/or relative liver weights at

≥ 10 mg/m 3 in males and ≥ 110 mg/m 3 in females, with the greatest increases at 110 and 250 mg/m 3 (21–PBBs and PBDEs 49

4. RELEVANCE TO PUBLIC HEALTH—PBDEs

44%). Centribobular hypertrophy similarly occurred in the liver at ≥ 10 mg/m 3 in both sexes (100% incidences in all groups except 4/5 females at 10 mg/m 3 ). Nasal effects included minimal to mild Goblet cell hyperplasia and/or hypertrophy at ≥ 1 mg/m 3 in males and ≥ 10 mg/m 3 in females. The nasal Goblet cell hyperplasia/hypertrophy occurred in nasal levels II and III in males at 1 mg/m 3 , and generally in nasal

levels II–VI in both sexes at ≥ 10 mg/m 3 . No changes were found in the remaining squamous, respiratory, and olfactory nasal epithelia, or other parts of the respiratory tract.

As detailed above, hepatocellular hypert rophy was found at concentrations at ≥ 3.7 mg/m 3 in the Great Lakes Chemical Corporation (1978) study and ≥ 10 mg/m 3 in the Great Lakes Chemical Corporation (2000) study. This liver effect was accompanied by some degenerative hepato cellular 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/m 3 in the Great Lakes Chemical Corporation (2000) study, but there were no nasal effects in the earlier study. Interpretati on of the significance and adversity of the liver and nasal alterations is complicated by the inconsistenc y of findings between the studies and other factors, including the small numbers of animals per study (fi ve/sex/level) and different methods of inhalation exposure (whole-body versus nose-onl y). Additionally, a well designed 13-week study (Great Lakes Chemical Corporation 2001a, 2001b) found (1) hepa tocellular hypertrophy, but no degenerative liver changes, at higher a minimum effect level (16 mg/m 3 ) than in both 14-day studies, and (2) no clearly exposure-related changes in nasal Goblet cells at concentrations below 202 mg/m 3 (study summarized in intermediate MRL section). Further, the adversity of the slight nasal Goblet cell changes, which were predominantly minimal in severity and possibly sugges tive of very slight nasal irritation in both the 14-day and 13-week studies, is unclear.

The available information indicates that there are insu fficient bases for considering the hepatic and nasal changes as adverse acute effects. More importantly, exposure to ≥ 16 mg/m 3 caused changes in serum levels of thyroid hormones (decreased T 3 , increased TSH) in the 13-week st udy. Thyroid hor mone 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, particular ly 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/m 3 has been derived for intermedia te-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/m 3 for changes in thyroid hormones in rats that were intermittently exposed to octaBDE for 13 weeks (Great Lakes Chemical Corporation 2001a, 2001 b). Calculation of the MRL is detailed below.

The inhalation database for intermed iate-duration exposure to PBDE s 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 produc t (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 (a ir only), 1.1, 16, or 202 mg/m 3 for 6 hours/day, 5 days/week, for

13 weeks. The mean MMADs in the low to hi gh exposure groups were 2.0, 2.7, and 2.8 μ m, and the corresponding mean GSDs were 3.37, 3.72, and 3.01. Clinical and physical signs, body weight, food consumption, and survival were ev aluated throughout the study. Ophthalmic, hematology (11 indices), serum chemistry (18 indices), and ser um thyroid hormone (TSH, total T 3 , and total T 4 ) evaluations were performed near the end of the exposure period. Urinalyses were not conducted. Comprehensive necropies, organ weight measurements, and histologi cal 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/m 3 and liver weight (absolut e and relative) at 202 mg/m 3 . Total incidences of centrilobular hepatocellular hype rtrophy in the 0, 1.1, 16, and 202 mg/m 3 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/m 3 , but showed no clear dose-related increasing tre nds 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/m 3 ; incidences in the 0, 1.1, 16, and 202 mg/m 3 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/m 3 (4/10, 0/10, 1/10, and 8/10, all minimal severity, not increased in females). Histological chan ges in the lungs included alveolar histiocytosis and chronic active inflammation that were only clearly increased in incidence at 202 mg/m 3 . Total incidences of alveolar histiocytosis at 0, 1.1, 16, and 202 mg/m 3 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 to tal 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 ofPBBs and PBDEs 51

4. RELEVANCE TO PUBLIC HEALTH—PBDEs

both lesion types tended to increase from minimal to mild/moderate at 202 mg/m 3 . Gross lung changes also occurred in both sexes at 202 mg/m 3 ; these included lung firmness and white discoloration and/or enlargement in the bronchial and/or mediastinal lym ph nodes. The lymph node effects correlated with the histological finding of granulomatous inflammati on. 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 T 4 )

at ≥ 16 mg/m 3 in both sexes, and increases in TSH at ≥ 16 mg/m 3 in males and 202 mg/m 3 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-i nduced hypothyroidism. There were no serum T 3 changes. Qualitative histological evaluations of step sections of ovaries showed an absence of corpora lutea in 3/10 females at 202 mg/m 3 , compared to 0/10 in the control and lower exposure groups. This 30% incidence was interpreted to be a treatment-relate d 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/m 3 ) and LOAEL (16 mg/m 3 ) for changes in serum levels of thyroid hormones, and abundant evidence for thyroid effects of PBDEs in or al studies, the effects on thyroid hormones are the most appropriate basis for estimation of an inte rmediate-duration inhalation MRL. The MRL of 0.006 mg/m 3 was derived by dividing the NOAEL HEC of 0.53 mg/m 3 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

NOAEL HEC was calculated using the following equations:

NOAEL ADJ = 1.1 mg/m 3 x 6 hours/24 hours x 5 days/7 days = 0.196 mg/m 3

NOAEL HEC = NOAEL ADJ x RDDR = 0.196 mg/m 3 x 2.7 = 0.53 mg/m 3

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 μ m with a mean GSD (sigma g) of 3.37, default human body we ight of 70 kg, and a default female F344 rat body weight of 0.18 kg. Additional information on the derivation of the intermedia te-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-durat ion 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 (C arlson 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/da y 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-Daw ley 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 impl antations, 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 (94

B 97% 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 ofPBBs 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. Comp rehensive histological examinations were performed, but limited to the control and high-dose groups. No hema tology, clinical chemistry, or urine indices, or thyroid hormone levels, were evaluated. There were no compound-related clin ical signs, deaths, body weight or food consumption changes, gross pathol ogy, or histopathology, indicating that 8,000 and 9,500 mg/kg/day are intermediate-duration NOAELs for sy stemic toxicity in rats and mice, respectively. As discussed above, the NOAEL for developmental toxi city 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 exposu re to decaBDE. Only one chronic study of high purity decaBDE has been conducted. In this study, a commercial decaBDE product (94

B 97% pure) was fed to F344 rats (50/sex/dose level) in dietar y 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). An imals 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 anim als 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 degenera tion 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 majo r 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 ob served. 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 a nd 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 acu te-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 T 4 hormone in fetal rats that were exposed to pentaB DE on days 4–20 of gestation (Zhou et al. 2002). The MRL was estimated by dividing the NOAEL by an uncerta inty 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 eff ects 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 T 4 and T 3 concentrations measured at each age point. The maternal exposure to pentaBDE caused reductions in serum total T 4 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 T 4 concentrations in the offspring w as age-dependent as values returned to control levels by PND 36. There were no exposure-related effects on serum total T 3 concentrations in the dams or offspring at any time, although T 3 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 T 4 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 T 4 levels in weanling (28-day-old) rats that were exposed to octaBDE fo r 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 fo r fetotoxicity in rats exposed to octaBDE for 10 days during gestation (Life Science Research Is rael 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 T 4 in rats and mice exposed to pentaBDE for 14 days (Fowles et al. 1994; Hallgren et al. 2001). Ad ditional 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 deba table. As discussed in the thyroid effects part of Section 4.2, mechanistic data indicate that human s are generally less sensitive than rodents becausePBBs and PBDEs 55

4. RELEVANCE TO PUBLIC HEALTH—PBDEs

(1) PBDEs likely affect thyroid function by displacing T 4 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 transporte r 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 T 4 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 T 4 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 Rese arch 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 T 3 and T 4 ), 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 evaluati on) 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 w eeks 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 week s 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. N onhepatic effects included decreases in serum T 4 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-dur ation studies that support selection of the 2 mg/kg/day minimal LOAEL include hepatic LOA ELs of 5 mg/kg/day for cytomegaly (with vacuolation and necrosisPBBs and PBDEs 56

4. RELEVANCE TO PUBLIC HEALTH—PBDEs

at higher doses) in rats exposed to octaBD E 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 T 4 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-Dawle y rats (25/sex/dose level) were fed a 77.4% pure commercial decaBDE mixture (containing 21.8% nonaBDE and 0.8% octaBD E) for approximately 2 years. Evaluations that included clinical signs, body weight, food consumption, hematology, clinical chemistry, urine indices, and comprehensive histol ogical examinations show ed 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 chro nic oral MRL based on this study would be 5 times higher than the 0.002 mg/kg/day intermediate MRL. A similar pattern was observe d for thyroid effects in the study used to derive the acute-duration oral MR L (Zhou et al. 2001) as summarized above. Due to the

insufficiencies of the chronic data for MRL derivatio n, 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 ove rall perspective on the toxicology of polybrominated biphenyls (PBBs) and polybrominated diphenyl ethers (P BDEs). It contains descriptions and evaluations of toxicological studies and epidemiological investig ations 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 hydrocar bons 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 struct ure 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 ox ygen bridge confers three-dimensi onal 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 approximat ely 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, octa brombiphenyl, 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 oc tabromobiphenyl PBB mixtures contained a large proportion (47.4–60.0%) of nonabromobiphenyl congen ers, whereas commercial decabromobiphenyls contained predominately (96.8%) decabromobiphenyl c ongener. The general na mes 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 decrease d 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 Administrati on (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 an alytical sensitivity (Dunckel 1975; Fries 1985a). The Michigan Department of Agriculture (MDA) subseque ntly 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 quarantin ing program established by the MDA in May 1974, about 30,000 dairy cattle, 2,000 swine, 400 sheep, an d over 2,000,000 chickens were found to contain PBBs at concentrations requiring their destruc tion (Dunckel 1975; Fries 1985a; Mercer et al. 1976).

Most of the information that is available on hea lth effects of PBBs in humans comes from studies of Michigan residents who ingested milk, meat, and e ggs 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 begi nning of testing and the establishment of quarantines, PBB-contaminated food products we re 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 contam ination 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 subs tantial increase in researc h activity regarding the health effects of PBBs in cattle, poultry, and la boratory animals. Compar ed to the FireMaster (commercial hexabromobiphenyl) PBBs, relatively li mited data are available on health effects of commercial mixtures of octabromobiphenyl and decabro mobiphenyl. 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 f eed (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 P BBs in livestock is insufficient to ascertain whether the people ingested PBBs or metabolic products of P BBs. However, based on available data discussed in Section 5.4.2.2, it is reasonable to assume that ma inly unchanged penta-, hexa -, and heptabromobiphenyl congeners in animal pr oducts 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 hei ghtened due to relatively recent evidence that some PBDE congeners have become ubiqu itously distributed in the environment and are present in tissues and breast milk of the general population at levels th at continue to increase. Three commercial PBDE mixtures have been produced: decabromodiphe nyl 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-, a nd hexaBDE congeners. Congeners with less than four bromine atoms are ge nerally 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). Pe nta- and octa-BDEs are apparently being phased out of production and use. Great Lakes Chemic al 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 informati on 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 pr ocedures for congener mixtures and commercially available standards for all congeners are lacking, and conge ner-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 othe r 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 transfor mation in the environment and/or differential biological metabolism and retention.

5.2 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE

To help public health professiona ls 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 (deat h, systemic, immunological, neurological, reproductive, developmental, genotoxic, and carcinog enic 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-obser ved-adverse-effect levels (NOAELs) or lowestobserved-adverse-effect levels (LOAELs) reflect the act ual 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 appro ach to warrant an attempt at distinguishing betweenPBBs and PBDEs 61

5. HEALTH EFFECTS

"less serious" and "serious" effects. The distinction be tween "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 NO AELs should also help in determining whether or not the effects vary with dose and/or duration, and pl ace into perspective the po ssible significance of these

effects to human health.

The significance of the exposure levels shown in th e 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 s ites 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. Es timates of levels posing minimal risk to humans (Minimal Risk Levels or MRLs) may be of inte rest 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 reliabl e and sufficient data exist to identify the target organ(s) of effect or the most sensitive health eff ect(s) for a specific duration within a given route of exposure. MRLs are based on noncancerous health e ffects only and do not consider carcinogenic effects. MRLs can be derived for acute, intermediate, a nd chronic duration exposures for inhalation and oral routes. Appropriate methodology does not exis t 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 techniqu es. Furthermore, ATSDR acknowledges additional uncertainties inherent in the application of the pro cedures 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 hypersensitiv ity 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 wo rkers 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 hea lth effects end points in each species and duration category for PBDEs are recorded in Table 5-1 and plotte d in Figure 5-1; due to the general lack of data regarding inhalation exposure to PBBs, no e quivalent table is presented for PBBs.

5.2.1.1 Death

Polybrominated Biphenyls. No studies were located regarding deat h in humans after inhalation exposure

to PBBs.

Nose-only exposure to the highest attainable dust con centration of octabromobiphenyl mixture for 4 hours (960 mg/m 3 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/m 3 concurrently with starch dust (6 mg/m 3 ) 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/m 3 , respectively, for 1 hour and observed for the following 14 days (IRDC 1974, 1975a, 1975b). Confidence in these studies is lim ited 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/m 3 for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978), ≤ 250 mg/m 3 for 6 hours/day, 5 days/week for 14 days (Great Lakes Chemical Corporation 2000), or ≤ 202 mg/m 3 for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2001a, 2001b).LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReferenceTable 5-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - InhalationChemical Form(mg/m3) (mg/m3) (mg/m3)Exposure/Duration/Frequency(Specific Route)Species(Strain)ACUTE EXPOSURESystemic13.73.7Resp20632424 (reversible rapid breathing)14 d8 h/d(CD)RatGreat Lakes Chemical Corporation 1978OctaBDE165165Cardio165165Gastro165165Hemato0.60.6Hepatic3.73.7 (hepatocytomegaly and focalhepatocellular degeneration)165165Renal165165Endocr165165Ocular165165Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - InhalationChemical Form(mg/m3) (mg/m3) (mg/m3)Exposure/Duration/Frequency(Specific Route)Species(Strain)211Resp211914d5 d/wk6 h/d(CD)RatGreat Lakes Chemical Corporation 2000OctaBDE11Hepatic1010250250Renal250250Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - InhalationChemical Form(mg/m3) (mg/m3) (mg/m3)Exposure/Duration/Frequency(Specific Route)Species(Strain)INTERMEDIATE EXPOSURESystemic31616Resp2116202202 (alveolar histiocytosis, chronicactive lung inflammation)13 wk5 d/wk6 h/d(CD)RatGreat Lakes Chemical Corporation 2001OctaBDE202202Cardio202202Gastro202202Hemato202202Musc/skel1.11.1Hepatic1616 (centrilobular hepatocellularhypertrophy)202202Renal1.11.1bEndocr1616 (decreased serum T4,increased serum TSH)202202Dermal202202Ocular202202Bd WtImmuno/ Lymphoret416162118202202 (grossly discolored andenlarged bronchial andmediastinal lymph nodesassociated with chronic activelung inflammation and alveolarhistiocytosis)13 wk5 d/wk6 h/d(CD)RatGreat Lakes Chemical Corporation 2001OctaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-1 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - InhalationChemical Form(mg/m3) (mg/m3) (mg/m3)Exposure/Duration/Frequency(Specific Route)Species(Strain)Reproductive51616 Fa 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 brominateddiphenyl ethers. The MRL was derived byconverting 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 uncertainty factor of 30 (3 forspecies to species extrapolation with dosimetric adjustments and 10 for human variability) and a modifying factor of 3 (for anincomplete data base).Bd Wt = body weight; Cardio = cardiovascular; d = day(s); Endocr = endocrine; F = female; gastro = gastrointestinal; hemato =hematological; hr = hour(s); LOAEL = lowestobserved adverse effect level; M = male; metab = metabolic; min = minute; mo = month(s); Musc/skel = muscular/skeletal; NOAEL =no observed adverse effect level; Resp =respiratory; T4 = thyroxine ; TSH = thyroid stimulating hormone; wk = week(s)2117202202 F (absence of corpora lutea inovaries)13 wk5 d/wk6 h/d(CD)RatGreat Lakes Chemical Corporation 2001OctaBDE0.11101001000 Respiratory1r1r2r Cardiovascular1r Gastrointestinal1r Hematological1r Hepatic1r1r2r2r Renal1r2r Endocrine1r Ocular1r Body Weight1r2rmg/m3Figure 5-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- InhalationAcute (≤14 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemic0.00010.0010.010.11101001000 Respiratory3r3r Cardiovascular3r Gastrointestinal3r Hematological3r Musculoskeletal3r Hepatic3r3r Renal3r Endocrine3r3r Dermal3r Ocular3r Body Weight3r Immuno/Lymphor4r4r Reproductive5r5rmg/m3Figure 5-1. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- Inhalation (Continued)Intermediate (15-364 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemicPBBs and PBDEs 69

5. HEALTH EFFECTS

5.2.1.2 Systemic Effects

Systemic effects that have been observed in huma ns 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 fe male rats that were exposed to a decabromobiphenyl dust mixture at 5,000 mg/m 3 concurrently with starch dust (6 mg/m 3 ) for 6 hours/day, 5 days/week for 4 weeks (Millischer et al. 1980). This effect was not observed at 500 mg/m 3 and lower concentrations or in air only-exposed controls, and there were no ch anges in pulmonary resistance and compliance in urethane-anesthetized rats, blood gases, and lung hist ology at any of the expos ure levels. Lung function and blood gases were not evaluated in starch-exposed c ontrols, 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 disso lved in corn oil), octaBDE dust, or decaBDE dust in very high concentrations of 200,000, 60,000, and 48,200 mg/m 3 , 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/m 3 as powdered dust for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corpor ation 1978). Increased respiration rate occurred at

≥ 23.9 mg/m 3 . 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/m 3 rats (other groups not examined) showed no changes inPBBs 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/m 3 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 go blet cell hyperplasia and/or hypertrophy at ≥ 1 mg/m 3 in males

and ≥ 10 mg/m 3 in females. The nasal goblet cell changes occurred in nasal levels II and III at 1 mg/m 3 and generally in levels II-VI at ≥ 10 mg/m 3 , and were considered indicative of very slight nasal irritation. No histopathology in the lungs or toxicolgical ly 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/m 3 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, occurre d in both sexes, and were only clearly induced at 202 mg/m 3 . Total incidences of alveolar histiocy tosis in the 0, 1.1, 16, and 202 mg/m 3 exposure groups were 3/10, 5/10, 5/10, and 10/10 in males, respectiv ely, and 0/10, 5/10, 2/10, and 10/10 in females, respectively. Respective total incidences of chroni c 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 fe males. 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/m 3 ; 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 granulomat ous 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/m 3 , but incidences were not clearly dose-related and there was essentially no increase in severity from minimal with increasing dose. Total inci dences of goblet cell hypertrophy in nasal level II in the 0, 1.1, 16, and 202 mg/m 3 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 fe males. Minimal severity goblet cell hypertrophy was also slightly increased in nasal level IV in males at 202 mg/m 3 (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/m 3 for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978), or ≤ 202 mg/m 3 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 gast rointestinal effects in humans or animals after inhalation exposure to PBBs.

Polybrominated Diphenyl Ethers. No studies were located regarding g astrointestinal effects in humans after inhalation exposure to PBDEs.

No histopathological changes were observed in the stom ach and lower gastrointestinal tract tissues of rats that were exposed to dusts of commercia l octaBDE products at levels of 174 mg/m 3 for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978), or ≤ 202 mg/m 3 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 fi ve female rats that were exposed to a decabromobiphenyl dust mixture at concentrations of 5 or 5,000 mg/m 3 concurrently with starch dust (6 mg/m 3 ) for 6 hours/day, 5 days/week for 4 weeks (Millischer et al. 1980). The evaluation included erythrocyte and leucocyte counts, differential leukocyte c ount, 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/m 3 of commercial octaBDE dust aerosol for 8 hours/day fo r 14 consecutive days (Great Lakes Chemical Corporation 1978). Evaluation of a limited number of indices (hemoglobin, hematocrit, total erythrocytePBBs 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 incr ease in leukocyte counts was considered to be an unusual response by the investigators, although it was w ithin the normal range for control rats in their laboratory. Comprehensive hematological assessmen ts showed no unusual changes in rats exposed to

commercial octaBDE as dust aerosol at concentrations of ≤ 202 mg/m 3 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 octa bromobiphenyl dust mixture at 960 mg/m 3 for 4 hours (timeweighted average, highest attainable concentrati on), and observed for 7 days (Waritz et al. 1977). Toxicity of octabromobiphenyl mixture vapor was inves tigated 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 μ g/m 3 , 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/m 3 concurrently with starch dust (6 mg/m 3 ) 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 ther efore may be an adaptive response because PBBs are hepatic inducers and cause cellular proliferation (see Sec tion 5.2.2.2 Hepatic Effects). No effects on liver weight or histology were observed at 5 mg/m 3

.

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 st udies of dusts of commercial octaBDE mixtures. In one study, rats were chamber-exposed to concentr ations of 0, 0.6, 3.7, 23.9, or 165.2 mg/m 3 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/m 3 and higher levels of exposure. AtPBBs and PBDEs 73

5. HEALTH EFFECTS

3.7 mg/m 3 , the liver lesions consisted of very slight to slight severity focal to multifocal cytoplasmic enlargement of the hepatocytes, accompanied by fo cal acidophilic degeneration of individual to small groups of cells. The liver lesions were similar at ≥ 24.4 mg/m 3 ), except that the hepatocyte enlargement was multifocal to diffuse in distribution and accompanie d by focal, small to large areas of hepatocellular necrosis of very slight to marked degree. In anothe r study, rats were nose-only exposed to 0, 1.0, 10, 110, or 250 mg/m 3 as dust aerosol for 6 hours/day, 5 days/week for 2 weeks (Great Lakes Chemical Corporation 2001a, 2001b). Hepatic effects include d increased mean absolute and/or relative liver weights at ≥ 10 mg/m 3 in males and ≥ 110 mg/m 3 in females, with the greatest increases at 110 and 250 mg/m 3 (21–44%). Centribobular hypertrophy similarly occurred in the liver at ≥ 10 mg/m 3 in both sexes (100% incidences except for 4/5 females at 10 mg/m 3 ).

Similar hepatic changes were found in a study of rats that were nose-only exposed to 0, 1.1, 16, or 202 mg/m 3 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/m 3 and increased liver weight (absolute and relative) at 202 mg/m 3 . Respective total incidences of cen trilobular hepatocellular hypertrophy (predominantly minimal to mild) in the 0, 1.1, 16, and 202 mg/m 3 groups were 1/10, 0/10, 3/10, and 10/10 in males, and 0/10, 0/10, 3/10, and 6/10 in fe males. Serum chemistry evaluations showed no clear effects of exposure. Serum cholesterol was significa ntly increased (39.8% less than controls, p<0.01) in 202 mg/m 3 females, but the magnitude of the elevation w as not considered toxicologically significant. Some other statistically significant serum chemistr y alterations (increased mean globulin and total protein, decreased albumin/globulin ratio) also occurred in the 202 mg/m 3 females, but were not considered exposure-related due to sm all 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 μ g/m 3 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 measureme nt 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 a nd five female rats that were exposed to a decabromobiphenyl dust mixture at concentrations ranging from 5 to 5,000 mg/m 3 concurrently with starch dust (6 mg/m 3 ) for 6 hours/day, 5 days/week for 4 weeks (Millischer et al. 1980). The analysis included pH, specific gravity, proteins, glucose, keto ne bodies, biliary pigments, urobilinogen, blood, and microscopic examination of sediment. A comprehe nsive 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; therefor e, 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 pr oducts at levels of 174 mg/m 3 for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978), ≤ 250 mg/m 3 for 6 hours/day, 5 days/week for 14 days (Great Lakes Chemical Corporation 2000), or ≤ 202 mg/m 3 for 6 hours/day, 5 days/week for 13 weeks (Great Lakes Chemical Corporation 2001a, 2001b). Urinal yses 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 onl y chemicals manufactured and who had volunteered for a comprehensive medical evaluation performed 3 m onths 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 μ U/ml normal range), low or borderline

low serum T 4 levels (4.4 versus 4.5–11.5 μ g/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 T 4 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 T 4 level (p=0.11) and antimicrosomal antibody titer (p=0.06) did not di ffer significantly from the controls. QuestioningPBBs and PBDEs 75

5. HEALTH EFFECTS

about exposure to 74 occupational toxicants showed th at 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 a nd 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 thyroi d enlargement, thyroid nodularity or hypothyroidism on physical examination, or had reported taking thyroi d 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 se rum 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 chemical s 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 d ecaBDE 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 (T 3 and free T 4 ) 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 hor mones and congeners.

Inhalation studies of commercial octaBDE dust in rats showed no histopathological changes in the thyroids, parathyroids, adrenals, or pituit ary following chamber exposure to 174 mg/m 3 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/m 3 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/m 3 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

T 4 ) at ≥ 16 mg/m 3 in both sexes, and increases in thyroid stimulating hormone (TSH) at ≥ 16 mg/m 3 in males and 202 mg/m 3 in females. The changes were usually st atistically significant (p<0.05 or p<0.01) compared to controls and were considered by the in vestigators to be consiste nt with chemical-induced hypothyroidism. There were no serum T 3 changes, thyroid-attributable clinical signs or body weightPBBs and PBDEs 76

5. HEALTH EFFECTS

effects, or gross or histopathological changes in the thyroid. The 1.1 mg/m 3 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 discu ssed 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 departme nts 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 (1 1.4–1,729 ppb) and 16.5 ppb (4–234 ppb); PBBs were not detected in serum of the control subjects (C handa et al. 1982). Physical examination confirmed

the occurrence of bromacne in 13% of PBB workers co mpared with no acne in the control group. No other studies were located regarding dermal effect s 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/m 3 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/m 3 concurrently with starch dust (6 mg/m 3 ) for 6 hours/day, 5 days/week for 4 weeks (Millischer et al. 1980). The seriousness of thisPBBs 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/m 3 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 ey e 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 concentr ations of 2,000, 2,000, and 48,200 mg/m 3 , 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/m 3 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/m 3 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 parame ters 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 le vels were not reported. Immunological analyses included determination of immunoglobu lin 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 respon ses (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 stimulat es 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 regardi ng immunological or lymphoreticular effects in humans after inhalation exposure to PBDEs.

No histopathological changes were observed in lymph no des or bone marrow from rats that were exposed to 174 mg/m 3 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/m 3 as powdered dust for 8 hours/day for 14 consecu tive days (Great Lakes Chemical Corporation 1978), or in the spleens (only lymphoreticular tissue reported) following nose-only exposure to

≤ 250 mg/m 3 as dust aerosol for 6 hours/day, 5 days/week fo r 14 days (Great Lakes Chemical Corporation 2000). Rats that were nose-only exposed to commerc ial octaBDE at levels of 1.1, 16, or 202 mg/m 3 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/m 3 . 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 cultu ral 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 eff ects 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/m 3 for 8 hours/day for 14 consecutive days (Great Lakes ChemicalPBBs and PBDEs 79

5. HEALTH EFFECTS

Corporation 1978), or ≤ 202 mg/m 3 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 (fo rebrain, 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 th e distribution of sperm counts, motility, or sperm morphology compared with a control group of 52 n onexposed 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 eff ects 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 t estes or ovaries from rats that were exposed to

commercial octaBDE at concentrations of ≤ 174 mg/m 3 as powdered dust for 8 hours/day for 14 consecutive days (Great Lakes Chemical Corporation 1978), or ≤ 250 mg/m 3 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 th at were nose-only exposed to 0, 1.1, 16, or 202 mg/m 3 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 qualitati ve evaluation of step sections of the ovary, was found in 3/10 females at 202 mg/m 3 , compared to 0/10 incidences in the control and both lower exposure groups. The investigators interpreted this 30% incr ease in incidence be treat ment-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 va gina, 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 effe cts 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 relia ble 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 su fficient to observe effects that develop slowly.

Except as noted below, acute-duration studies administer ed 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/da y 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 aLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReferenceTable 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)ACUTE EXPOSUREDeath128610001000 (18/18 died)2 wk5d/wk1x/d(GO)(Fischer344/N)RatGupta and Moore 1979(FF-1)219130130 F (63% lethality)14 dad lib(F)(Balb/c)MouseFraker 1980;Fraker and Aust 1978(BP-6)Systemic3286286 MEndocr55once(GO)(Sprague-Dawley)RatAllen-Rowlands et al. 1981(NS)411bEndocr5833 M (decreased thyroid plasma T4hormone)10 d1x/d(GO)(SpragueDawley)RatAllen-Rowlands et al. 1981(NS)5400400 FBd Wt50800800 F (unknown percent maternalweight loss)onceGd 6-14(GO)(Wistar)RatBeaudoin 1977(BP-6)6Hepatic99500500 M (increased hepaticphospholipids and serumcholesterol)once18 mo observ(GO)(Sherman)RatBernert et al. 1983(FF-1)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)7Hepatic28910001000 (hepatocytic swelling, fattyinfiltration, multinucleation,necrosis, and cytolysis)2 wk5 d/wk1x/d(GO)(Fischer344/N)RatGupta and Moore 1979(FF-1)Renal10001000 (darkened kidneys)Endocr10001000 (darkened adrenal glands)Bd Wt10001000 (unknown percent weight loss,emaciation)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)83030Resp3102 wk5 d/wk1x/d(GO)(Fischer344/N)RatGupta et al. 1981(FF-1)3030Cardio3030Gastro3030Hemato3030Musc/skel0.30.3Hepatic33 (dose-related hepatocyteenlargement and single-cellnecrosis)3030Renal3030Endocr3030Dermal3030Ocular3030Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)910001000Resp206once2-14 mo observ(GO)(Sherman)RatKimbrough et al. 1978(FF-1)10001000Cardio10001000GastroHepatic10001000 (vacuolation, necrosis, andfibrosis, porphyria,multinucleation)10001000Renal10001000Endocr10001000Bd Wt10Hepatic184500500 M (vacuolation of hepatocytes)once15wk observ(GO)(Sherman)RatKimbrough et al. 1980(FF-1)500500 MBd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)11200200 FResp187once18-22mo observ(GO)(Sherman)RatKimbrough et al. 1981(FF-1)200200 FCardioHepatic200200 F (porphyrin accumulation)200200 FRenal200200 FEndocr200200 FDermal200200 FBd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)1210001000 FResp198once23mo observ(GO)(Sherman)RatKimbrough et al. 1981(FF-1)10001000 FCardio10001000 FGastro10001000 FMusc/skelHepatic10001000 F (hepatomegaly, hepatocyteenlargement and vacuolation,porphyrin accumulation)10001000 FRenal10001000 FEndocr10001000 FOcularBd Wt10001000 F (12% decreased body weightgain)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)13Hepatic9310001000 M (fatty changes in centrilobularhepatocytes)once28d observ(GO)(SpragueDawley)RatLee et al. 1975a(OBB)10001000 MBd Wt14Hepatic9430003000 M (fatty changes in centrilobularhepatocytes)2 d1x/d(GO)(SpragueDawley)RatLee et al. 1975a(OBB)10001000 MBd Wt1520002000 FBd Wt165once(GO)(Sprague-Dawley,Spartan)RatNorris et al. 1975b(OBB)16Hepatic17955 M (hepatomegaly and fattychanges in weanlings)10 dad lib(F)(Fischer 344)RatRaber and Carter 1986(BP-6)55 MBd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)170.660.66 MHepatic1026.536.53 M (hyperplasia and fatty changes)2 wkad lib(F)(SpragueDawley)RatWaritz et al. 1977; Lee et al. 1975b(OBB)7171 MRenal7171 MEndocr7171 MBd Wt183636 FHepatic1392 wkad lib(F)(SwissWebster)MouseCagen et al. 1977(BP-6)19Hepatic47130130 F (focal areas of coagulativenecrosis)11 dad lib(F)(Swiss/ IRC)MouseCorbett et al. 1975(BP-6)20Bd Wt96130130 M (30% decreased body weight)4-14 dad lib(F)(Swiss/ IRC)MouseCorbett et al. 1978(BP-6)21Bd Wt21130130 F (23% weight loss)14 dad lib(F)(Balb/c)MouseFraker 1980;Fraker and Aust 1978(BP-6)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)223030Resp3062 wk5 d/wk1x/d(GO)(B6C3F1)MouseGupta et al. 1981(FF-1)3030Cardio3030Gastro3030Hemato3030Musc/skel0.30.3Hepatic33 (dose-related increase inincidence of hepatocyteenlargement and single-cellnecrosis)3030Renal3030Endocr3030Ocular3030Bd Wt3030OtherLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)Immuno/ Lymphoret2328710001000 (atrophy of thymus; necrosis ofsplenic lymphoblasts)10 d5 d/wk1x/d(GO)(Fisher344/N)RatGupta and Moore 1979(FF-1)2420130130 F (suppressed antibody-mediatedresponse to SRBC, thymicatrophy)14 dad lib(F)(Balb/c)MouseFraker 1980;Fraker and Aust 1978(BP-6)Reproductive2519610001000 F (9% increased incidence ofuterine polyps)once23mo observ(GO)(Sherman)RatKimbrough et al. 1981(FF-1)262121 F736363 F (29% reduction in success ofpregnancy)9 dGd 6-151x/d(F)(C57BL)MouseWelsch and Morgan 1985(HBB)Developmental27404049200200 (9.1-31.4% resorptions)onceGd 6-14(GO)(Wistar)RatBeaudoin 1977(BP-6)2855435050 (12% decrease in fetal bodyweight)14 dGd 7-20(F)(SpragueDawley)RatCorbett et al. 1975(BP-6)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)29160200200 (increased mortality and liverneoplasms in offspring)Gd 7-14 1x/d(GO)(Sherman)RatGroce and Kimbrough 1984(FF-1)304242.942.9 (12-20% decreased mean bodyweight in treated pups atpost-parturition day 60)9 dGd 7-151x/d(GO)(NS)RatHarris et al. 1978(BP-6)311000100039910 dGd 6-151x/d(GO)(SpragueDawley, Iffacredo)RatMillischer et al. 1980(DBB)329.19.12648686 (increased incidence of extraribs)10 dGd 6-15ad lib(F)(ChR-CD)RatWaritz et al. 1977(OBB)3355455050 (cleft palate)12 dGd 7-18(F)(Swiss/ IRC)MouseCorbett et al. 1975(BP-6)Cancer34161200200 (CEL: hepatocellular carcinomain offspring)Gd 7-14 1x/d(GO)(Sherman)RatGroce and Kimbrough 1984(FF-1)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)3519710001000 F (CEL: hepatocellular carcinoma)once(GO)(Sherman)RatKimbrough et al. 1981(FF-1)INTERMEDIATE EXPOSUREDeath36290149149 M (90-day LD50)6565 F (90-day LD50)c4.5 wk5 d/wk22 d(GO)(Fischer344/N)RatGupta and Moore 1979(FF-1)371440.30.3 M (decreased mean survival time)25 wk5 d/wk1x/d(GO)(Fischer344/N)RatNTP 1983(FF-1)381501010 F (decreased mean survival time)25 wk5 d/wk1x/d(GO)(B6C3F1)MouseNTP 1983(FF-1)3926144 M (4/6 died)30 dad lib(F)(NS)Gn PigSleight and Sanger 1976(NS)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)4012822 F (7/8 died)45 dad lib(F)(NS)Gn PigVos and van Genderen 1973(BP-6)412500.470.47 M (LD50)c0.610.61 F (LD50)313 dad lib(F)(NS)MinkAulerich and Ringer 1979(FF-1)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)Systemic42Gastro2851818 F (hyperplastic gastroenteritis)137 dad lib(F)(Rhesus)MonkeyAllen et al. 1978(FF-1)Hemato1818 F (decreased RBC, PCV, andWBC)Hepatic1818 F (enlarged hepatocytes,hyperplasia of bile ductepithelium, increased SGPT,decreased serum cholesterol)Renal1818 F (hyperplasia of bladderepithelium)Endocr1818 F (adrenal hemorrhage)Dermal1818 F (edema, atrophy, andsquamous metaplasia ofsebaceous glands)Bd Wt1818 F (27% body weight loss)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)43Cardio2760.730.73 M (enlarged heart at necropsy)25-50 wkad lib(F)(Rhesus)MonkeyAllen et al. 1978; Lambrecht et al. 1978(FF-1)Gastro0.730.73 M (proliferation of mucosal cells,chronic inflammatory cells,severe ulcerative colitis)Hemato0.730.73 M (decreased PCV and totalserum protein)Hepatic0.730.73 (enlarged hepatocytes withincreased lipid droplets,increased SGPT, decreasedserum cholesterol, hyperplasiaof bile duct epithelium)Dermal0.730.73 (edema and alopecia,keratinization of hair folliclesand sebaceous glands)Bd Wt0.730.73 (34% weight loss in adult male,0% weight gain in juvenile)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)441010 MResp5430 dad lib(F)(SpragueDawley)RatAkoso et al. 1982a(BP-6)1010 MCardio1010 MGastro1010 MHemato1010 MMusc/skelHepatic0.10.1 M (hepatocyte swelling,vacuolation)1010 MRenal1010 MEndocr1010 MDermal1010 MOcular1010 MBd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)45Endocr980.050.05 M (altered thyroid follicularultrastructure)30 dad lib(F)(SpragueDawley)RatAkoso et al. 1982b(BP-6)46Endocr5611 M (decreased serum thyroidhormone T4)20 d1x/d(GO)(SpragueDawley)RatAllen-Rowlands et al. 1981(NS)47Endocr320.450.45 F (decreased thyroid serum T3and T4 hormones)7 moad lib(F)(SpragueDawley)RatByrne et al. 1987(BP-6)480.050.05 FEndocr350.250.25 F (decreased adrenal serumcorticosterone B, DHEA andDHS hormones)5-7 moad lib(F)(SpragueDawley)RatByrne et al. 1988(BP-6)4966 MEndocr3420 d1x/d(GO)(SpragueDawley)RatCastracane et al. 1982(NS)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)5022 MResp34228 dad lib(F)(SpragueDawley)RatChu et al. 1980(BP-6)22 MCardio22 MGastro22 MHematoHepatic22 M (increased liver weight,increased liver microsomalenzymes, fatty degeneration ofliver)22 MRenalEndocr22 M (reduction of follicular size andcolloid density and exfoliation ofepithelium in thyroid)22 MDermal22 MBd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)51Hepatic2210.50.5 M (bile duct hyperplasia)82 dad lib(F)(SpragueDawley)RatDarjono et al. 1983(BP-6)Ocular55 M (xerophthalmia)55 MBd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)5210001000Cardio2944.5 wk5 d/wk1x/d(GO)(Fischer344/N)RatGupta and Moore 1979(FF-1)10001000GastroHemato3030 (decreased hemoglobin, PCV,and platelet count)Hepatic3030 (hepatocyte enlargement, fattyinfiltration and multinucleation,porphyrin accumulation)Renal3030 (dilation of Bowman's capsulewith serous fluid)Endocr3030 (unspecified altered thyroidhistology)Bd Wt3030 (19% decreased body weightgain)100100 (emaciation)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)533030Resp29830 d5 d/wk1x/d(GO)(Fischer344/N)RatGupta et al. 1981(FF-1)3030Cardio3030Gastro3030Hemato3030Musc/skel0.30.3Hepatic33 (increased liver weight,hepatocyte swelling, andnecrosis)3030Renal33Bd Wt3030 (significant decrease in bodyweight)54Endocr310.250.25 M (colloid droplets, abnormalmicrovilli and other changes inthyroid follicle ultrastructure)5 wkad lib(F)(Holtz- man)RatKasza et al. 1978a(BP-6)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)550.250.25 MHepatic972.52.5 M (hepatocyte hypertrophy anddegeneration)5 wkad lib(F)(Holtz- man)RatKasza et al. 1978a(BP-6)56Hemato121010 F (increased white blood cellcount)6 mo5 d/wk(GO)(Fischer 344)RatLuster et al. 1980(FF-1)11 FBd Wt33 F (15% decreased weight gain)57Hepatic17855 F (enlarged and vacuolatedhepatocytes, focal necrosis)3 moad lib(F)(SpragueDawley)RatMcCormack et al. 1978(BP-6)Renal55 F (degenerative changes inglomeruli)Bd Wt55 F (10% decreased body weightgain)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)58100100Gastro39713 wk(F)(Sprague-Dawley CFY)RatMillischer et al. 1980(DBB)100100Hemato2525Hepatic100100 (11% increased liver weight,hepatocyte hypertrophy andvacuolization, slightly increasedliver lipids)100100Renal100100Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)59800800 MCardio17330 dad lib(F)(SpragueDawley)RatNorris et al. 1975b(OBB)8080 MHemato800800 M (decreased PCV and RBCcounts)Hepatic88 M (enlargement and vacuolation)Renal88 M (hyaline degenerative changes)Endocr88 M (thyroid hyperplasia)800800 MBd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)6011Cardio1778 moad lib(F)(SpragueDawley)RatNorris et al. 1975b(OBB)11Hemato11Hepatic11Renal11Endocr11Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)611010Resp14925 wk5 d/wk1x/d(GO)(Fischer344/N)RatNTP 1983(FF-1)1010Cardio0.30.3Gastro11 (gastric ulcers)0.10.1Hemato0.30.3 (decreased hemoglobin, MCH,PCV, and MCV)1010Musc/skel0.10.1Hepatic0.30.3 (lipid accumulation; increasedatypical foci; porphyrinaccumulation)0.30.3Renal11 (chronic progressivenephropathy)0.10.1Endocr0.30.3 (decreased serum thyroid T4hormone)1010Ocular1010Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)622.52.5 FHepatic597 moad lib(F)(SpragueDawley)RatSepkovic and Byrne 1984(OBB)2.52.5 FEndocr632.52.5 FEndocr607 moad lib(F)(SpragueDawley)RatSepkovic and Byrne 1984(HBB)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)645050 MResp25930 dad lib(F)(SpragueDawley)RatSleight and Sanger 1976(NS)5050 MCardio5050 MGastro5050 MHemato5050 MMusc/skelHepatic11 M (hepatocyte vacuolation)1010 MRenal5050 M (unquantified, but significantlyincreased BUN)5050 MEndocr1010 MBd Wt5050 M (16% decreased weight gain)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)6555 MResp15830 dad lib(F)(SpragueDawley)RatSleight et al. 1978(BP-6)55 MCardio55 MGastro55 MHemato0.50.5 MHepatic55 M (hepatocyte swelling andvacuolation)55 MRenal0.50.5 MEndocr55 M (hyperplasia of thyroid follicularepithelium)0.50.5 MBd Wt55 M (27-36% reduced body weightgain)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)660.660.66 MHepatic1076.536.53 M (hyperplasia and progressivelipid changes)4 wkad lib(F)(SpragueDawley)RatWaritz et al. 1977; Lee et al. 1975b(OBB)7171 MRenal7171 MEndocr7171 MBd Wt671313 FBd Wt2330 dad lib(F)(Balb/c)MouseFraker 1980;Fraker and Aust 1978(BP-6)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)683030Resp30230 d5 d/wk1x/d(GO)(B6C3F1)MouseGupta et al. 1981(FF-1)3030Cardio3030GastroHemato3030 F (decreased PCV)3030Musc/skel0.30.3Hepatic33 (hepatocyte enlargement andnecrosis)3030Renal3030Ocular33Bd Wt3030 M (significant decrease in weightgain)3030OtherLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)6921.721.7 MResp666 wkad lib(F)(Balb/ cBYJ)MouseLoose et al. 1981(FF-1)21.721.7 MCardio0.650.65 MHepatic21.721.7 M (hepatocellular necrosis andvacuolation)21.721.7 MRenal0.650.65 MEndocr0.650.65 MBd Wt21.721.7 M (33% reduction in body weight)701010Hemato146 mo5 d/wk(GO)(B6C3F1)MouseLuster et al. 1980(FF-1)Endocr1010 (increased adrenal weight gain)1010Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)711010Resp15525 wk5 d/wk1x/d(GO)(B6C3F1)MouseNTP 1983(FF-1)1010Cardio1010Gastro0.10.1Hemato0.30.3 (decreased erythrocyte countand MCV)1010Musc/skel0.10.1Hepatic0.30.3 (increased liver weight, SGOTand porphyrin accumulation)1010Renal1010Endocr1010Ocular33 MBd Wt1010 M (25% decreased body weightgain)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)722020 MResp26230 dad lib(F)(NS)Gn PigSleight and Sanger 1976(NS)Hepatic0.040.04 M (vacuolation and fatty changes)0.40.4 MBd Wt44 M (severe weight loss prior todeath)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)7388Cardio18216 wkad lib(F)(NS)PigKu et al. 1978(NS)11Gastro88 (gross hyperplasia glandularstomach)88HematoHepatic11 (LDH increased)88RenalEndocr88 (increased adrenal weight)11Dermal88 (dermatosis)Bd Wt11 (12.9% reduced weight gain andfood intake)740.1250.125 FHepatic101.251.25 F (fatty changes and necrosis)12 wkGwk 8-ppwk 4ad lib(F)(NS)PigWerner and Sleight 1981(BP-6)1.251.25 FEndocr2.52.5 F (significant decrease in thyroidserum T3 and T4 hormones)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)752.42.4Cardio255313 dad lib(F)(NS)MinkAulerich and Ringer 1979;Ringer andAulerich 1981(FF-1)Hepatic0.240.24 F (48% increased relative liverweight, fatty infiltration)2.42.4RenalBd Wt0.390.39 F (14% decreased prebreedingbody weight gain)1.861.86 F (up to 19% mean body weightloss prior to death)Immuno/ Lymphoret760.030.03 M2433 M (decreased lymphocyticresponse to mitogenstimulation; decrease inabsolute and relative thymusweight)5 wk5 d/wk(GO)(Fischer 344)RatLuster et al. 1978(FF-1)7711 F1133 F (decreased lymphoproliferativeresponses and decreaseddelayed hypersensitivityresponses)6 mo5 d/wk1x/d(GO)(Fischer 344)RatLuster et al. 1980(FF-1)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)780.130.13 F221.31.3 F (reduced antibody mediatedresponse to SRBC and 21%reduction in thymus weight)30 dad lib(F)(Balb/c)MouseFraker 1980;Fraker and Aust 1978(BP-6)790.650.65 M6521.721.7 M (increased lethality due toendotoxin challenge)6 wkad lib(F)(Balb/ cBYJ)MouseLoose et al. 1981(FF-1)8033131010 (increased lethality due toinfection with L monocytogenes;decreased response to mitogenstimulation)6 mo5 d/wk1x/d(GO)(B6C3F1)MouseLuster et al. 1980(FF-1)811290.40.4 F (reduced antitoxin titersfollowing toxoid challenge)44 F (thymic atrophy and folliculardepletion in spleen)45 dad lib(F)(NS)Gn PigVos and van Genderen 1973(BP-6)821.251.25 F1102.52.5 F (reduced lymphocyte responseto mitogen stimulation)12 wkGwk 8- ppwk 4ad lib(F)(NS)PigHoward et al. 1980(BP-6)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)Neurological83741010 (decreased limb strength)6 mo5 d/wk1x/d(GO)(NS)RatCabe and Tilson 1978(FF-1)8433 M16266 M (decreased motor activity)4 wk5 d/wk1x/d(G)(SpragueDawleyHoltz- man)RatGeller et al. 1979(FF-1)850.20.2100222 (delayed acquisition oflocomotion and reduced openfield activity in offspring)40dGd 6-Ppd 24(F)(SpragueDawley)RatHenck et al. 1994(BP-6)868833 (decreased motor activity, gripstrength, and startleresponsiveness)6 mo5 d/wk(GO)(Fischer344/N)RatTilson and Cabe 1979(FF-1)87893030 (decreased open field motoractivity and grip strength)4 wk5 d/wk1x/d(GO)(Fischer344/N)RatTilson and Cabe 1979(FF-1)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)Reproductive882750.730.73 M (hypoactive seminiferoustubules)25-50 wkad lib(F)(Rhesus)MonkeyAllen et al. 1978; Lambrecht et al. 1978(FF-1)892460.0120.012 F (increased menstrual cycleduration in 4/7; implantationbleeding in 2/7)6 moad lib(F)(Rhesus)MonkeyLambrecht et al. 1978; Allen et al. 1978;1979(FF-1)9014.314.3 F13728.628.6 F (no implantations in 2/5 rats)15 dGd 0-148x(GO)(Wistar)RatBeaudoin 1979(BP-6)913030 Mc10001000 F293100100 M (squamous metaplasia,hyperplasia, and necrosis inepithelium of ductus deferens)4-5 wk5 d/wk22 d(GO)(Fischer344/N)RatGupta and Moore 1979(FF-1)9255 F14242 dGd 8-ppd 28ad lib(F)(SpragueDawley)RatMcCormack et al. 1981(BP-6)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)9330303014-5 wk5 d/wk1x/d(GO)(B6C3F1)MouseGupta et al. 1981(FF-1)940.240.242540.390.39 F (10% reduction in body weight)313 dad lib(F)(NS)MinkAulerich and Ringer 1979(FF-1)Developmental952.92.913614.314.3 (increased resorptions)15 dGd 0-148x(GO)(Wistar)RatBeaudoin 1979(BP-6)96551094 wk5 d/wk(G)(Holtz- manSpragueDawley)RatGeller et al. 1985(FF-1)971080.20.2 (deficits in learning behavior inoffspring, 6 months afterprenatal and lactationalexposure)40 dGd 6- ppd24ad lib(F)(SpragueDawley)RatHenck and Rech 1986(BP-6)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)9810010.20.2 M (reduced crown-rump length)40dGd 6-Ppd 24(F)(SpragueDawley)RatHenck et al. 1994(BP-6)991410.50.5 (increased liver weight,hepatocyte vacuolation,decreased hepatic vitamin Acontent in F1 but not F2)55 (decreased pup survival duringlactation in F1)42 dGd 8-ppd 28ad lib(F)(SpragueDawley)RatMcCormack et al. 1981(BP-6)1006755 (20% decrease in pup bodyweight gain, 50% decreasedhepatic vitamin A, 256-285%decreased urinary uro- andcoproporphyrins in pups)42-126 dGd 8-ppd 28-112ad lib(F)(SpragueDawley)RatMcCormack et al. 1982a(BP-6)101612.52.5 (decreased body weight,increased relative liver weight,and decreased serum T4 inoffspring)37 dGd 0-ppd 15ad lib(F)(SpragueDawley)RatMeserve et al. 1992(BP-6)1023580.50.5 (hepatic vacuolization andaltered foci in pups)77 dGd 0-ppd 56ad lib(F)(Fischer344/N)RatNTP 1992, Chhabra et al. 1993(FF-1)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)103221533 (decreased hematocrit inoffspring)1010 (early postnatal death; nodetails provided)42 dGd 0- weaning1x/2d(GO)(B6C3F1)MouseLuster et al. 1980(FF-1)1043711.51.5 (hepatic cytomegaly and alteredfoci in pups)77 dGd 0- ppd 56ad lib(F)(B6C3F1)MouseNTP 1992, Chhabra et al. 1993(FF-1)1051733 (performance deficits inoffspring in a learned task)Gd 0-ppd 211x/2d(GO)(C57B1/6)MouseTilson 1992(FF-1)1060.1250.12591.251.25 (increased relative liver weight,decreased serum thyroidhormone levels, and slightthyroid hyperplasia in offspring)12 wkGwk 8- ppwk 4ad lib(F)(NS)PigWerner and Sleight 1981(BP-6)1072530.1550.155 (decreased birth and 4-weekweights in kits)313 d(F)(NS)MinkAulerich and Ringer 1979;Ringer andAulerich 1981(FF-1)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)Cancer108191100100 (CEL: hepatocellular carcinoma)4 mo12 x(GO)(Sherman)RatKimbrough et al. 1981(FF-1)10914833 (CEL: hepatocellular carcinoma)25 wk5 d/wk1x/d(GO)(Fischer344/N)RatNTP 1983(FF-1)1101541010 (CEL: hepatocellularcarcinomas)25 wk5 d/wk1x/d(GO)(B6C3F1)MouseNTP 1983(FF-1)1113701.51.5 (CEL: hepatocullular adenomaand carcinoma in offspring)77 dGd 0-ppd 56ad lib(F)(B6C3F1)MouseNTP 1992, Chhabra et al. 1993(FF-1)CHRONIC EXPOSUREDeath1123470.50.5 M (18% decreased survival)115 wksGd 0-ppd 56 (weaning)104 wks post-weaningad lib(F)(Fischer344/N)RatNTP 1992, Chhabra et al. 1993(FF-1)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)1133601.31.3 F (44% decreased survival)116 wksGd 0-ppd 56 (weaning)105wks post-weaningad lib(F)(B6C3F1)MouseNTP 1992, Chhabra et al. 1993(FF-1)Systemic1140.0120.012Hemato24766 wkad lib(F)(Rhesus)MonkeyLambrecht et al. 1978(FF-1)Bd Wt0.0120.012 (7.4% weight loss)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)1151.51.5Resp352115 wksGd 0-ppd 56;104 wks post-weaning(F)(Fischer344/N)RatNTP 1992, Chhabra et al. 1993(FF-1)1.51.5Cardio0.50.5Gastro1.51.5 M (forestomach hyperplasia,inflammation, ulceration)0.50.5Hemato1.51.5 F (mild anemia)1.51.5Musc/skelHepatic0.50.5 (hypertrophy, vacuolation,altered foci, increased serumcholesterol, decreased serumtriglycerides)1.51.5Renal1.51.5Endocr1.51.5DermalBd Wt0.50.5 (11-18% decreased final bodyweight)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)1163.93.9Resp365116 wksGd 0-ppd 56 (weaning)105 wks post-weaningad lib(F)(B6C3F1)MouseNTP 1992, Chhabra et al. 1993(FF-1)3.93.9Cardio3.93.9Gastro3.93.9Musc/skelHepatic1.31.3 (hypertrophy, vacuolization,single-cell necrosis, altered foci,bile duct hyperplasia)1.31.3Renal3.93.9 (increased chronic nephropathy)Endocr1.31.3 (thyroid follicular cellhyperplasia)3.93.9Dermal3.93.9Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)Immuno/ Lymphoret1170.50.5 M3481.51.5 M (splenic fibrosis)115 wksGd 0-ppd 56 (weaning)104 wks post-weaningad lib(F)(Fischer344/N)RatNTP 1992, Chhabra et al. 1993(FF-1)1181.31.33613.93.9 (increased splenichematopoiesis)116 wksGd 0-ppd 56 (weaning)105 wks post-weaningad lib(F)(B6C3F1)MouseNTP 1992, Chhabra et al. 1993(FF-1)Reproductive1191.51.5 M0.50.5 Fc3501.51.5 F (cystic endometrial hyperplasia)115 wksGd 0-ppd-56 (weaning)104 wks post-weaningad lib(F)(Fischer344/N)RatNTP 1992, Chhabra et al. 1993(FF-1)1203.93.9363116 wksGd 0-ppd 56 (weaning)105 wks post-weaningad lib(F)(B6C3F1)MouseNTP 1992, Chhabra et al. 1993(FF-1)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-2 Levels of Significant Exposure to Polybrominated Biphenyls - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)Developmental1212450.0120.012 (1/7 fetuses were aborted,1/7fetuses stillborn, 12%decreased birth weight and 22%decreased postnatal weight gainin 4/7 survivors)359-469 dad lib(F)(Rhesus)MonkeyLambrecht et al. 1978; Allen et al. 1978;Allen et al. 1979(FF-1)Cancer1223511.51.5 (CEL: leukemia)0.50.5 (CEL: hepatocellular adenomaand carcinoma)115 wksGd 0- ppd 56104 wks post-weaning(F)(Fischer344/N)RatNTP 1992, Chhabra et al. 1993(FF-1)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 NOAELby an uncertainty factor of 100 (10 for extrapolationfrom 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 levels of effect for themost 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 biphenyl; DW = drinkingwater; endocr = endocrine; (F)= feed; F = female; FF-1 = FireMaster FF-1; (G) = gavage; gastro = gastrointestinal; Gd = gestation day; Gn Pig - Guinea Pig; (GO) = gavage in oil;Gwk = gestation week; HBB = hexa-brominated biphenyl; hemato = hematological; hr = hour(s); LD50 = lethal dose; 50% kill; LOAEL= lowest observed adverse effect level; M =male; min = minute(s); mo = month(s); Musc/skel = musculoskeletal; NOAEL = no observed adverse effect level; NS = not specified; OBB octa-brominated biphenyl; observ =observation; ppd = post partum day; ppwk = post partum week; Resp = respiratory; T4 = thyroxine; wk = week(s); x = time(s)3643.93.9 (thyroid follicular cell adenoma)1.31.3 (CEL: hepatocellular adenomaand carcinoma)116 wksGd 0- ppd 56105wks post-weaning(F)(B6C3F1)MouseNTP 1992, Chhabra et al. 1993(FF-1)0.010.1110100100010000 Death2m1r Respiratory22m8r9r11r12r Cardiovascular22m8r9r11r12r Gastrointestinal22m8r9r12r Hematological22m8r Musculoskeletal22m8r12r Hepatic18m19m22m22m6r7r8r8r9r10r11r12r13r14r16r17r17r Renal22m7r8r9r11rmg/kg/dayFigure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- OralAcute (≤14 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancer12r17rSystemic0.010.1110100100010000 Endocrine22m3r4r4r7r8r9r11r12r17r Dermal8r11r Ocular22m8r12r Body Weight20m21m22m5r5r7r8r9r10r11r12r13r14r15r16r17r Other22m Immuno/Lymphor24m23r Reproductive26m26m25r Developmental33m33m27r27r28r28r29r30r31r32r32r Cancer *34r35rmg/kg/dayFigure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continued)Acute (≤14 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemic*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.0.010.1110100100010000 Death39g40g41n38m36r37r Respiratory72g68m69m71m44r50r53r61r64r65r Cardiovascular75n43k68m69m71m44r50r52r53r59r60r61r64r65r73p Gastrointestinal42k43k68m71m44r50r52r53r58r61r61r64r65r73p73p Hematological42k43k68m70m71m71m44r50r52r53r56r58r59r59r60r61r61r64r65r73p Musculoskeletal68m71m44r53r61r64rmg/kg/dayFigure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continued)Intermediate (15-364 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemic0.010.1110100100010000 Hepatic55r55r57r58r58r59r60r61r61r62r64r65r65r66r66r73p74p74p Renal75n42k68m69m71m44r50r52r53r57r58r59r60r61r61r64r64r65r66r73pmg/kg/dayFigure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continued)Intermediate (15-364 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancer

Endocrine42k69m70m71m44r45r46r47r48r48r49r50r52r54r59r60r61r61r62r63r64r65r65r66r73p74p74p72g75n42k43k68m68m69m69m71m71m44r50r51r52r53r53rSystemic0.010.1110100100010000 Body Weight72g72g75n75n42k43k67m68m68m69m69m70m71m71m44r50r51r52r52r53r53r56r56r57r58r59r60r61r64r64r65r65r66r73p Other68mmg/kg/dayFigure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continued)Intermediate (15-364 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancer

Dermal42k43k44r50r73p73p Ocular68m71m44r51r61rSystemic0.010.1110100100010000 Developmental95r96r97r98r99r99r100r101r102r106p106p Cancer *110m111m108r109rmg/kg/dayFigure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continued)Intermediate (15-364 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancer

Neurological83r84r84r85r85r86r87r Reproductive94n94n88k89k93m90r90r91r91r92r107n103m103m103m104m105m95r

Immuno/Lymphor81g81g78m78m79m79m80m80m76r76r77r77r82p82p*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.0.010.1110 Death113m112r Respiratory116m115r Cardiovascular116m115r Gastrointestinal116m115r115r Hematological114k115r115r Musculoskeletal116m115r Hepatic116m115r Renal116m116m115r Endocrine116m115r Dermal116m115r Body Weight114k116m115r Immuno/Lymphor118m118m117r117r Reproductive120m119r119r Developmental121k Cancer *123m123m122r122rmg/kg/dayFigure 5-2. Levels of Significant Exposure to Polybrominated Biphenyls- Oral (Continued)Chronic (≥365 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemic*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.LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReferenceTable 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg) (mg/kg) (mg/kg)Exposure/Duration/Frequency(Specific Route)Species(Strain)ACUTE EXPOSUREDeath1210662006200 (44-day LD50)once(GO)(Wistar)RatBritish Industrial Biological ResearchAssociation 1977PentaBDE2209050005000 (4/5 died)once(GO)SpartanRatIRCD 1975bPentaBDE3210750005000 (14-day LD50)once(GO)(Sprague-Dawley)RatPharmakon Research International Inc.1984PentaBDESystemic41010Bd Wt2071100100 (30% reduced maternal bodyweight gain)10 dGd 6-15(GO)(CD)RatArgus Research Laboratories 1985aPentaBDE576.676.6 MHepatic204214 d1x/d(GO)(SpragueDawley)RatCarlson, 1980bOctaBDE656.456.4 MHepatic204314 d1x/d(GO)(SpragueDawley)RatCarlson, 1980bPentaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)7Endocr21091818 (reduced serum T4)14 d1x/d(GO)(SpragueDawley)RatDarnerud and Sinjari 1996PentaBDE8Hepatic20681818 (reduced liver vitamin A)14 d1x/d(GO)(SpragueDawley)RatHallgren et al. 2001PentaBDEEndocr1818 (reduced serum T4)950005000Bd Wt2089once(GO)SpartanRatIRCD 1975aOctaBDE10500500Bd Wt2091once(GO)SpartanRatIRCD 1975bPentaBDE112525Bd Wt207510 dGd 6-15(GO)(CD)RatLife Science Research Israel Ltd. (1987)OctaBDE122525Bd Wt20735050 (40% reduced maternal bodyweight gain)10 dGd 6-15(GO)(CD)RatWIL Research Laboratories 1986OctaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)131010Endocr20813030 (reduced serum T4)4 d1x/d(GO)(LongEvans)RatZhou et al. 2001PentaBDE1433Endocr20821010 (reduced serum T4)4 d1x/d(GO)(LongEvans)RatZhou et al. 2001OctaBDE15Endocr21101818 (reduced serum T4)14 d1x/d(GO)(C57BL/6N)MouseDarnerud and Sinjari 1996PentaBDE16500500Hepatic2059once(GO)C57BL/6JMouseFowles et al. 1994PentaBDE100100Endocr500500 (reduced serum T4)177272Hepatic206114d1x/d(GO)(C57BL/6N)MouseFowles et al. 1994PentaBDEEndocr1818 (reduced serum T4)7272Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)181818Hepatic20693636 (reduced liver vitamin A)14 d1x/d(GO)(C57BL/6N)MouseHallgren et al. 2001PentaBDEEndocr1818 (reduced serum T4)Immuno/ Lymphoret193636206614 d1x/d(GO)(SpragueDawley)RatDarnerud and Thuvander 1998PentaBDE20181820673636 (reduced in vitro production ofIgG in mitogen- stimulatedsplenocytes)14 d1x/d(GO)(C57BL/6N)MouseDarnerud and Thuvander 1998PentaBDE215005002060once(GO)(C57BL/6N)MouseFowles et al. 1994PentaBDE22363620627272 (reduced antibody response tosheep red blood cells,decreased thymus weight)14d1x/d(GO)(C57BL/6N)MouseFowles et al. 1994PentaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)Developmental23200200207010 dGd 6-15(GO)(CD)RatArgus Research Laboratories 1985aPentaBDE24101020582525 (increased resorptions andreduced fetal body weight)10 dGd 6-15(GO)(CD)RatArgus Research Laboratories 1985bOctaBDE252.52.520741010 (minimal increasedpost-implantation loss)10 dGd 6-15(GO)(CD)RatLife Science Research Israel Ltd. (1987)OctaBDE26252520725050 (reduced fetal weight andincreased skeletal variationsassociated with maternal tox)10 dGd 6-15(GO)(CD)RatWIL Research Laboratories 1986OctaBDE2711b21111010 (reduced serum T4 in fetuses)14 dGd 6-Gd 20(GO)(LongEvans)RatZhou et al. 2002PentaBDE2855 F20451515 F (delayed ossification ofsternebrae with decreasedmaternal weight gain)13 dGd 7-19(GO)(NewZealand)RabbitBreslin et al. 1989OctaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)INTERMEDIATE EXPOSURESystemic291.771.77 MHepatic203990 d(GO)(Sprague-Dawley)RatCarlson 1980aPentaBDE3014.114.1 MHepatic204090 d(GO)(Sprague-Dawley)RatCarlson 1980aPentaBDE312.42.4 MHepatic203790 d(GO)(Sprague-Dawley)RatCarlson, 1980aOctaBDE3219.219.2 MHepatic203890 d(GO)(Sprague-Dawley)RatCarlson, 1980aOctaBDE33Hepatic205399 (increased liver weight andenlarged parenchymal cells)28 d(F)(CD)RatIRDC 1976PentaBDE9090Renal9090Endocr9090Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)34Hepatic205599 (increased liver weight andenlarged parenchymal cells)28 d(F)(CD)RatIRDC 1976OctaBDE9090Renal9090Endocr9090Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)35750750 FResp207713 wk(F)(CD)RatIRDC 1977OctaBDE750750 FCardio750750 FGastro7070 FHemato750750 F (reduced erythrocytes,hematocrit and hemoglobin)Hepatic55 M (cytomegaly with vacuolationand necrosis at higher doses)5050 MRenal600600 M (minimal increase in tubulardegenerative changes)77 FEndocr5050 M (increased thyroid weight withfollicular epithelial changes athigher doses)750750 FDermal750750 FOcular7070 FBd Wt600600 M (12% reduced body weight gain)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)36100100Resp204890d(F)(Sprague-Dawley)RatWIL Research Laboratories 1984PentaBDE100100Cardio100100Gastro100100Hemato100100Musc/skelHepatic22 (minimal LOAEL forhypertrophy, mild degeneration,and slight necrosis)c100100Renal22Endocr1010 (reduced serum T4)100100Dermal100100Ocular1010Bd Wt100100 (reduced weight gain)LOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)371010Endocr20873030 (reduced maternal serum T4)36 dGd 6-Pnd 21(GO)(LongEvans)RatZhou et al. 2002PentaBDEImmuno/ Lymphoret38750750 F207813 wk(F)(CD)RatIRDC 1977OctaBDE39100100204990d(F)(Sprague-Dawley)RatWIL Research Laboratories 1984PentaBDEReproductive40600600 Md750750 F208013 wk(F)(CD)RatIRDC 1977OctaBDE41100100 M100100 F205190d(F)(Sprague-Dawley)RatWIL Research Laboratories 1984PentaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-3 Levels of Significant Exposure to Lower Brominated Diphenyl Ethers - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)Developmental4211a 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 NOAELby an uncertainty factor of 30 (10 for extrapolationfrom 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 factor of 300 (3 forconverting 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 levels of effect for themost 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) = gavage; gastro =gastrointestinal; (GO) = gavage in oil; hemato = hematological; hr = hour(s); LD50 = lethal dose; 50% kill; LOAEL = lowest observed adverse effect level;min = minute(s); M = male;Musc/skel = musculoskeletal; NOAEL = no observed adverse effect level; NS = not specified; pnd = post natal day; Resp = respiratory; T4 = thyroxine ; wk = week(s); x = time(s)20861010 (reduced serum T4 in fetusesand offspring on Gd 20 and Pnd4 and 14)36 dGd 6-Pnd 21(GO)(LongEvans)RatZhou et al. 2002PentaBDE0.010.1110100100010000 Death1r2r3r Hepatic16m17m18m18m5r6r8r Endocrine15m16m16m17m18m7r8r13r13r14r14r Body Weight17m4r4r9r10r11r12r12r Immuno/Lymphor20m20m21m22m22m19r Developmental23r24r24r25r25r26r26r27r27r28h28hmg/kg/dayFigure 5-3. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- OralAcute (≤14 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemic0.0010.010.11101001000 Respiratory35r36r Cardiovascular35r36r Gastrointestinal35r36r Hematological35r35r36r Musculoskeletal36r Hepatic29r30r31r32r33r34r35r36r Renal33r34r35r35r36r Endocrine33r34r35r35r36r36r37r37rmg/kg/dayFigure 5-3. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- Oral (Continued)Intermediate (15-364 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemic0.0010.010.11101001000 Developmental42r42rmg/kg/dayFigure 5-3. Levels of Significant Exposure to Lower Brominated Diphenyl Ethers- Oral (Continued)Intermediate (15-364 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancer

Dermal35r36r Ocular35r36r Body Weight33r34r35r35r36r36r Immuno/Lymphor38r39r Reproductive40r41rSystemicLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReferenceTable 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)ACUTE EXPOSURESystemic195.995.9 MHepatic204114 d1x/d(GO)(SpragueDawley)RatCarlson 1980bDecaBDE250005000Bd Wt2085once(GO)SpartanRatIRCD 1974DecaBDE31600016000Bd Wt200214d1x/d(F)(Fischer- 344)RatNTP 198694-97% decaBDE4100100Endocr20834 d1x/d(GO)(LongEvans)RatZhou et al. 2001DecaBDE51900019000Bd Wt200514d1x/d(F)(B6C3F1)MouseNTP 198694-97% decaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)INTERMEDIATE EXPOSURESystemic69090Hepatic205728 d(F)(CD)RatIRDC 1976DecaBDE9090Renal9090Endocr9090Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)780008000Resp200813 wk(F)(Fischer- 344)RatNTP 198694-97% decaBDE80008000Cardio80008000Gastro80008000Hemato80008000Musc/skel80008000Hepatic80008000Renal80008000Endocr80008000Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)895009500Resp201313 wk(F)(B6C3F1)MouseNTP 198694-97% decaBDE95009500Cardio95009500Gastro95009500Hemato95009500Musc/skel95009500Hepatic95009500Renal95009500Endocr95009500Bd WtImmuno/ Lymphoret980008000200913 wk(F)(Fischer- 344)RatNTP 198694-97% decaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)1095009500201413 wk(F)(B6C3F1)MouseNTP 198694-97% decaBDEReproductive1180008000201113 wk(F)(Fischer- 344)RatNTP 198694-97% decaBDE1295009500201613 wk(F)(B6C3F1)MouseNTP 198694-97% decaBDEDevelopmental1310001000 Fb211220 dGd 0-19(GO)(SpragueDawley)RatHardy et al. 200299% decaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)CHRONIC EXPOSURESystemic1411Resp20302 yr(F)(Sprague-Dawley)RatKociba et al. 1975; Norris et al. 1975b77% decaBDE, 22% nonaBDE11Cardio11Gastro11Hemato11Musc/skel11Hepatic11Renal11Endocr11Ocular11Bd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)1525502550 FResp2018103 wk(F)(Fischer- 344)RatNTP 198694-97% decaBDE25502550 FCardio11201120 MGastro22402240 M25502550 FHemato25502550 FMusc/skel12001200 MHepatic11201120 M22502250 F (pre-neoplastic nodules)25502550 FRenal25502550 FEndocr25502550 FBd WtLOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)1677807780 FResp2024103 wk(F)(B6C3F1)MouseNTP 198694-97% decaBDE77807780 FCardio37603760 FGastro77807780 F (ulcers)77807780 FHemato77807780 FMusc/skelHepatic32003200 M (centrilobular hypertrophy andgranulomas)77807780 FRenalEndocr32003200 M (follicular cell hyperplasia)77807780 FBd WtImmuno/ Lymphoret171120312 yr(F)(Sprague-Dawley)RatKociba et al. 1975; Norris et al. 1975b77% decaBDE, 22% nonaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)18201922402240 M (splenic fibrosis and lymphoidhyperplasia)12001200 F (splenic hematopoiesis)103 wk(F)(Fischer- 344)RatNTP 198694-97% decaBDE1977807780 F2025103 wk(F)(B6C3F1)MouseNTP 198694-97% decaBDEReproductive2011 M11 F20332 yr(F)(Sprague-Dawley)RatKociba et al. 1975; Norris et al. 1975b77% decaBDE, 22% nonaBDE2122402240 M25502550 F2021103 wk(F)(Fischer- 344)RatNTP 198694-97% decaBDELOAELLess SeriousNOAEL SeriousaSystemKey tofigureReference(continued)Table 5-4 Levels of Significant Exposure to Decabromodiphenyl Ether - OralChemical Form(mg/kg/day) (mg/kg/day) (mg/kg/day)Exposure/Duration/Frequency(Specific Route)Species(Strain)2266506650 M77807780 F2027103 wk(F)(B6C3F1)MouseNTP 198694-97% decaBDECancer23202211201120 M (CEL: liver neoplastic nodules)103 wk(F)(Fischer- 344)RatNTP 198694-97% decaBDE24a 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 wasderived by dividing the 1,000 mg/kg/day NOAELby 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 levels of effect for themost 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 = gestation day; (G) =gavage; gastro = gastrointestinal; (GO) = gavage in oil; hemato = hematological; hr = hour(s); LOAEL = lowest observed adverseeffect level; M = male; min = minute(s); Musc/skel =musculoskeletal; NOAEL = no observed adverse effect level; NS = not specified; Resp = respiratory; wk = week(s); x = time(s); yr = year(s)202832003200 M (CEL: hepatocellular adenomasand carcinomas)103 wk(F)(B6C3F1)MouseNTP 198694-97% decaBDE10100100010000100000 Hepatic1r Endocrine4r Body Weight5m2r3rmg/kg/dayFigure 5-4. Levels of Significant Exposure to Decabromodiphenyl Ether - OralAcute (≤14 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemic10100100010000 Respiratory8m7r Cardiovascular8m7r Gastrointestinal8m7r Hematological8m7r Musculoskeletal8m7r Hepatic8m6r7r Renal8m6r7r Endocrine8m6r7r Body Weight8m6r7r Immuno/Lymphor10m9r Reproductive12m11r Developmental13rmg/kg/dayFigure 5-4. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (Continued)Intermediate (15-364 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemic0.1110100100010000 Respiratory16m14r15r Cardiovascular16m14r15r Gastrointestinal16m16m14r15r15r Hematological16m14r15r Musculoskeletal16m14r15r Hepatic16m14r15r Renal16m14r15r Endocrine16m14r15r Ocular14r Body Weight16m14r15r Immuno/Lymphor19m17r18r Reproductive22m20r21r Cancer *24m23rmg/kg/dayFigure 5-4. Levels of Significant Exposure to Decabromodiphenyl Ether - Oral (Continued)Chronic (≥365 days)c-Catd-Dogr-Ratp-Pigq-Cow-Humansk-Monkeym-Mouseh-Rabbita-Sheepf-Ferretj-Pigeone-Gerbils-Hamsterg-Guinea Pign-Minko-OtherCancer Effect Level-AnimalsLOAEL, More Serious-AnimalsLOAEL, Less Serious-AnimalsNOAEL - AnimalsCancer Effect Level-HumansLOAEL, More Serious-HumansLOAEL, Less Serious-HumansNOAEL - HumansLD50/LC50Minimal Risk Levelfor effectsother thanCancerSystemic*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 hu nchback 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 pos ttreatment 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 th an rats and mice was not located.

In intermediate-duration studies with rats, no deaths we re 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% mo rtality in male and female rats, respectively; the average times to death were 46.7 and 60.3 days, r espectively (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 LD 50 in rats observed for ≈ 60 days posttreatment (i.e., 90-day LD 50 ) was 149 and 65 mg/kg/day for male and female rats, respectiv ely (Gupta and Moore 1979). This study did not specifically address the cause of death, but emacia ted 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 w eeks 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 (C hhabra et al. 1993; NTP 1992). The decreased survival appeared to be related to increased incidences of mononuclear cell leukemia. No deaths were observed inPBBs 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 m onths 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 informati on 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 P BBs 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 LD 50 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 (6 3–294 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 onl y 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 sim ilarly 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 se xes at dosages that decreased survival.

No deaths occurred in two swine th at 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 characteristicPBBs 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 LD 50 value and reliable LOAEL values for death in each species in the acute- and intermediateduration categories are recorded in Tabl e 5-2 and plotted in Figure 5-2.

Polybrominated Diphenyl Ethers. No deaths occurred in rats that we re 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 da ys (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 th at 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). Inte rmediate-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 LD 50 values of 5,000 and 6,200 mg/kg were determined for pentaBDE (Saytex 115 and DE-71, respectively) in rats that were observe d for 14 days (British Industrial Biological Research Association 1977; Pharmakon Research International In c. 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 toPBBs 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 LD 50 and LOAEL values for death in the acute-dur ation 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 followi ng 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 toxi cants 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 lung s, trachea, or nasal turbinates following gavagePBBs 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 na sal 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 pu rulent 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 ti ssues 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 a nd 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 car diovascular 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 studi es. 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 posttreatme nt (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 mixt ure for 30 days (Sleight and Sanger 1976). Rats exposed to ≤ 5 mg/kg/day FireMaster BP-6 for 30 days ex hibited no exposure-related changes in blood pressure, but histology or other cardiovascular end poin ts 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 we re reported in mink that died following exposure to

≤ 2.4 mg/kg/day FireMaster FF-1 for 313 days (Aul erich 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 histolog y 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 a nd 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 th at 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 sh owed 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 (N TP 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 y ears (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 Hea lth (Landrigan et al. 1979). In a medical history survey conducted by the Environmental Science Labo ratory of the Mount Sinai School of Medicine, no statistically significant difference was observed be tween the prevalence rate s 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 effect s 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) (Gupt a 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 histopatho logical 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 hyperplasticPBBs 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 ga vage treatment period, although similar changes (forestomach hyperplasia, inflammation, and ulcera tion) 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 ev aluation was performed in this study, but the liver is the only tissue specifically mentioned as having b een examined. Due to th e total number of tissues examined (21) and route of exposure, it is probable th at 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 de tails were not reported, and hi stology 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 in flammatory 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 gastroenter itis and, in the low-dose monkey (that died of "severe gastrointestinal ch anges"), severe ulcerativ e 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). Gastrointes tinal 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 betwee n days 33 and 66 because of impending death due to poor health showed edema and hemorrhage of the colon and r ectum mucosa, although histology was normal in the esophagus, rumen, omasum, and reticulum. No histologi cal 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 g astrointestinal 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 a nd 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 tr act 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 possi ble 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 s howed no exposure-relate d 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 significa ntly 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 administ ered the same dosages of FireMaster FF-1 ( ≤ 30 mg/kg/day) for 30 days, longer postexposure (up to 90 days) eval uation 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 return ed to control levels by 60-days post-dosing. No consistent hematological changes we re 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 atPBBs 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 we re 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 m onkeys 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 decrease d 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). Simila r 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 beca use the animals at this dose were sacrificed between days 33 and 66 because of impe nding death due to poor health.

Studies of hematologic effects of octabromobiphe nyl 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 intermed iate-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 hematol ogical effects in rats exposed by diet to

≤ 1.0 mg/kg/day of a 77.4% decaBDE mixture (cont aining 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 co mplaints in two epidemiological studies of groups of Michigan residents who were likely to have ing ested 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 sy mptoms 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 fo r symptoms of musculoskeletal effects.

There are no pathology data indicating that PBBs produ ce 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 th e 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 sacrificePBBs 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 (Koc iba 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 f ound 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 be tween 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 predominatel y 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 resi dents 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 s ubjects displayed mildly enlarged livers, which were associated with elevations in serum liver enzym es (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 w ith 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 micr osomal enzymes in exposed Michigan residents (Lambert et al. 1990).

Hepatic effects of PBBs are docum ented 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 proliferati on of the smooth endoplasmic reticulum, microsomal enzyme induction, increased serum levels of liv er-associated enzymes indicative of possible hepatocellular damage, liver enlarg ement, hepatocyte vacuolation and fat deposition, fibrosis, and

necrosis. PBBs also cause alterations in levels of c holesterol and other lipids in liver and serum, levels of vitamin A in liver and urine, and levels of porphyri ns 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 metabo lism 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 patte rn 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 micr osomal 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 singl e 1,000 mg/kg dose and observed for 2–23 months posttreatment or a lethal dose of 1,000 mg/kg/day fo r 2 weeks developed enlarged livers with fatty and necrotic changes leading to fibrosis (Gupta a nd Moore 1979; Kimbrough et al. 1978b, 1981). Lower single doses of FireMaster FF-1 caused vacuolation a nd 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 da ys (Raber and Carter 1986) caused hepatocyte enlargement and some fatty and single-cell necrotic changes in weanling and young rats. A limitedPBBs and PBDEs 176

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amount of data suggest that octabromobiphenyl mixture-i nduced 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 obser ved 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 c onsistent effects on disposition of injected ouabain or indocyanine green, indicating that hepatic function w as 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 dege nerative 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 th e only chronic study, incidences of hepatocellular hypertrophy, cytoplasmic vacuolation, atypical foci, a nd 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 peri natal and adult exposure resulted in increased incidences of oval cell hyperplasia at 0.5 mg/kg/day and hypertrophy, cyt oplasmic vacuolation, and bile duct fibrosis at 1.5 mg/kg/day. Studies of octabromobiphenyl mixt ure in rats have show n hepatic effects (e.g., hypertrophy and hyperplasia of centrilobular cells, vac uolation, 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 sl ightly increased lipid, did not occur at dosages <100 mg/kg/day (Millischer et al. 1980). Informati on on hepatic effects of octabromobiphenyl mixture and decabromobiphenyl in species othe r than the rat was not located.PBBs and PBDEs 177

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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 hyp erplasia also developed (NTP 1992). Dosages ≥ 3 mg/kg/day for 4–6 weeks, but not 0.3 mg/kg/day, also induced hepatocy te 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 Sl eight 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 in filtration (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 duc t 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 l actic 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 hepatocyt es, sinusoidal dilation, and scattered areas of early fatty degeneration. In general, the hepatic effect s 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 effect s of acute-duration oral exposure to PBDEs in

animals. Intermediate- and chronic-duration studies in r odents indicate that the liver is a target of toxicity of PBDEs. As discussed below, repeated dietar y exposure to commercial PBDE products typicallyPBBs and PBDEs 178

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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 pentaB DE 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 cau sed 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 ra ts was characterized by a near total occlusion of a major heptatic blood vessel by a dense fibrin coa gulum. A NOAEL was not identified in the rats or mice. The only other chronic study of decaBDE f ound 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. 1975 b). 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 histopath ological changes such as he patocellular 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 expo sed 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, cha nge in hepatocytic cytoplasm to a finely granular, homogeneous type, and cytopl asmic 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

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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 aff ected 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 in creased incidence of degeneration and necrosis of individual liver parenchymal cells at 24 weeks postex posure; 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, respec tively, for 90 days (Hallgren et al. 2001).

Hepatic microsomal enzyme induc tion is a well-documented effect for the pentaBDE and octaBDE commercial products, but has not been observed fo r decaBDE (Carlson 1980b; Zhou et al. 2001). Microsomal enzyme activity was induced in rats e xposed by gavage to doses as low as 0.6 mg/kg/day ofoctaBDE 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-i nduced increases in liver weight and microsomal enzyme activity (e.g., increased EPN detoxification, p- nitroanisole demethyla tion, uridinediphosphateglucuronyltransferase [UDPGT] activity, and benzo[a] pyrene hydroxylase activity); exposure to decaBDE only increased liver weight (Carlson 1980b). D ecaBDE also had no effect on hepatic UDPGT, ethoxyresorufin-o-deethylase (EROD), or pentoxyreso rufin-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

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Hepatic EROD, PROD, and methoxyresorufin-o-deethy lase (MROD) were incre ased 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 matern ally-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 correla tions were found between serum PBB

levels and serum levels of blood ur ea nitrogen (BUN) or creatinine in a study of residents of quarantined Michigan farms after the 1973 PBB contamination epis ode (Kreiss et al. 1982). No other studies were located with information pertinent to renal e ffects in humans after oral exposure to PBBs.

Studies with animals have shown some renal effect s 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 wa s 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 ex posure to FireMaster BP-6 for 30 days produced no PBB-related alterations in urinalysis indices or B UN at 5 mg/kg/day (highest tested dose) or kidneyPBBs and PBDEs 181

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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, inac tive, or had been largely replaced by scar tissue), although relative kidney weight, BUN, and re nal function tests (clearance of inulin, p -aminohippurate, or fractional sodium excretion) were nor mal (McCormack et al. 1978). Also, in vitro accumulation of

p -aminohippurate and N-methylnicotinamide, and a mmoniagenesis 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 gava ge exposure to a higher FireMaster FF-1 dose of 30 mg/kg/day for 4.5 weeks caused dilation of Bowman's capsule wi th 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 hist ology, 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 cha nges in urinalysis values or kidney histology (Sleight and Sanger 1976). In studies with octabrom obiphenyl 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 findi ng 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 pe rformed at this and lower dosages in this study, but the liver is the only tissue specifically mentione d as having been examined. Due to the total numberPBBs and PBDEs 182

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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 inci dence 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 (M oorhead 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 tubul es, degenerative changes in the tubular epithelium, and congestion with scatte red microscopic hemorrhages in the medulla. The renal effects in cows appear to be more severe th an those generally observed in other species at lethal doses. No urinalysis alterations or changes in kidne y 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 eff ects that are mainly attributable to octaBDE. Renal effects induced by octaBDE included noninfla mmatory 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

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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 ra ts 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 endo crine effects in humans after oral exposure to PBBs.

Thyroid effects have been observed in animals treat ed with PBBs by gavage or diet in acute-, intermediate-, and chronic-duration studies. Charact eristic changes include decreases in serum levels of serum thyroxine (T 4 ) and serum triiodothyronine (T 3 ) hormones, thyroid enlargement, and effects in the follicular cells including reduced size, hyperplasia wi th columnar appearance and papillary projections, and accumulation of colloid droplets. In the only acute study that inv estigated thyroid end points more sensitive than histology, rats administered an uns pecified PBB mixture for 10 days showed serum T 4 levels (T 3 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 T 4 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 T 4 , 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 131 I uptake and incorporation into thyroglobulin in rats (AllenRowlands et al. 1981). No thyroid histological altera tions 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 thyroi d 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 speciesPBBs and PBDEs 184

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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 T 3 or T 4 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 FireM aster 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 131 I and deceased incorporation of 131 I into monoiodotyrosine (MIT) at 6 mg/kg/day (Allen-Rowlands et al. 1981). No effects on incorporation of 131 I into diiodotyrosine (DIT), T 3 , or T 4 were observed. Serum T 4 levels were

also reduced at ≥ 1 mg/kg/day in rats exposed for 20 days an d evaluated after being placed on restricted

food intake for ≥ 2 months following treatment (Allen-Rowla nds et al. 1981). Rats administered 2.5 mg/kg/day of an unspecified hexabromobiphenyl mi xture for 7 months showed no significant changes

in serum T 3 , but serum T 4 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 ch anges 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 evaluate d 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 incl uded 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 T 3 and T 4 at 5 mg/kg/day. Chronic exposure to

≤ 1.5 mg/kg/day FireMaster FF-1 for up to 104 w eeks 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 T 3 , but serum T 4 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

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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 discrepa ncy in the octabromobiphenyl mixture NOAELs of

≤ 71 mg/kg/day and LOAELs of ≥ 8 mg/kg/day is not apparent, particul arly since treatment durations were similar, methods of treatment (diet) and animal st rain 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 Fire Master 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 fo r 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 adre nal 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, dehydroepiandroste rone (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 corticoste rone levels and adrenal weight di d 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 ob served in rats in other intermedia te-duration studies with FireMaster BP-6 or FireMaster FF-1 (Akoso et al. 1982b; NTP 1 983; 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 th e only rat study of octabromobiphenyl mixture that examined the adrenal gland, 2.5 mg/kg/day for 7 mont hs produced no changes in relative adrenal weight; histology or serum corticosteroids were not evalua ted (Sepkovic and Byrne 1984). Intermediate- or

chronic-duration studies with FireMaster FF-1 in mi ce 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 organoha logen compounds (non-PBDEs), as we ll as hormone levels (free and

total T 3 and T 4 , thyroid stimulating hormone [TSH], free testosterone, follicle-stimulating hormone, lutenizing hormone, and prolactin), were analyzed in 110 men who consumed varying amounts of fattyPBBs and PBDEs 186

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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 (r 2 =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 incr eased in male B6C3F1 mi ce that were exposed to

≥ 94% pure commercial decaBDE for 103 weeks (NTP 198 6). Incidences of the lesion were 2/50 (4%), 10/50 (20%), and 19/50 (38%) in the 0, 3,200, and 6,6 50 mg/kg/day dose groups of this study. Slight

increases in follicular cell tumors that were cons idered 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 hyperplasi a 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 in termediate- 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 indivi dual 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 T 4 , occurs by a mechanism that is not clearly relevant to humans (see Section 5.5.2 Mechanisms of Toxicity).

Thyroid hormone levels were determ ined 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 pentaB DE (DE-71) in doses of 0.3, 1, 3, 10, 30, 100, orPBBs and PBDEs 187

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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 T 4 , total T 3 , and TSH. DecaBDE caused no changes in levels of any of the thyroid hormones. Octa BDE induced a dose-related reduction in serum T 4 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 T 3 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 T 4 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 T 3 levels were significantly reduced at ≥ 100 mg/kg/day with a maximum reduction of 30% at 300 mg/kg/day. Neither octaBDE nor pentaBDE cau sed exposure-related changes in serum TSH concentrations. Benchmark dose (BMD) analysis of the octaBDE da ta found that the BMD and BMDL (95% lower confidence limit on the BMD) resulting in a 20% reduction in thyroid hormones (LED 20 ) were 9.25 and 5.29 mg/kg/day, respectively, for serum T 4 and 53.38 and 11.98 mg/kg/day, respectively, for serum T 3 . For pentaBDE, the respective BMD and BMDL resulti ng in 20% reduced hormone levels were 12.74 and 6.95 mg/kg/day for serum T 4 and 32.94 and 8.56 mg/kg/day for serum T 3 .

Serum levels of thyroid hormones were not evalua ted 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, r espectively. 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 co mpound-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 ep ithelium 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 cons idered 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-Evan s 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 T 4 and T 3 were evaluated in dams on GD 20 and PND 22 and offspring were evaluated on GD 20 and PNDsPBBs and PBDEs 188

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4, 14, 36, and 90. Serum T 4 was significantly (p<0.05) reduced co mpared 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 T 4 concentrations in the offspring returned to control levels by PND 36. The NOAEL for the effect on serum T 4 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 T 4 levels were reported to be 2.36 and 0.94 mg/kg/day, respectively. There were no exposurerelated changes in serum T 3 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 T 4 levels additionally occurred in adult offspring of rats that were exposed to commercial pentaBDE mixture DE-71 during gestation and lactation in two studies incomplete ly reported as abstracts (Taylor et al. 2002, 2003). In one study, rats were or ally 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 T 4 levels; specific data were not reported, although the BMD and BMDL for reduced T 4 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 horm ones in offspring showed dose-related decreases in

T 4 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 T 4 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 unde r 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 incomplete ly reported as abstracts. End points in both studies included pubertal development (time to pre putial separation in males and vaginal opening in females), body and reproductive tissue weights, thyroid function (serum levels of T 4 , T 3 , and TSH), and liver microsomal enzyme induction (EROD, PROD, a nd UDPGT activity). Both sexes showed delays in reproductive development that were accompanied by changes in thyroid horm one levels and hepatic metabolism. In the study with males, Wistar rats were gavaged with 0, 3, 30, or 60 mg/kg/day doses inPBBs and PBDEs 189

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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 incr eases in the age of preputial separation at 30 and 60 mg/kg/day (2.0- and 2.3-day delay, respectively). Seminal vesical and ventra l prostate weights were reduced at 60 mg/kg/day with no significant changes in testis or epididymal weights or serum testosterone levels. Thyroid hormone measuremen ts showed reductions in serum T 4 at 30 and 60 mg/kg/day (74 and 81% lower than controls) after 5 days of exposure, serum T 4 at 3, 30, and 60 mg/kg/day (20, 80, and 86% lower) after 31 days, and serum T 3 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 11 3% higher than controls) after 31 days. Hepatic effects included increased relative liver weight a nd 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 T 4 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 le vels were related to in duction 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 Spra gue-Dawley rats included serum T 3 and T 4 levels (TSH not measured) and thyroid histology (WIL Research Laboratories 1984). Effects observed in both sexes included significantly reduced plasma T 4 levels at ≥ 10 mg/kg/day and increased follicular cell hyperplasia at 100 mg/kg/day. Incidences of follicular cell hyperplasi a 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 ch aracterized 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). Inciden ces of slight or moderate hyperplasia in the 0, 9, or 90 mg/kg/day dose groups of this st udy were 0/5, 1/5, and 3/5 males, respectively; no increases were seen

in females.

Serum total T 4 concentrations were significantly reduced in female C57BL/6J mice that were given a single gavage dose of commercial pentaBDE mixtur e DE-71 as low as 0.8 mg/kg (lowest tested dose)PBBs and PBDEs 190

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(Fowles et al. 1994). Reductions in serum T 4 were found in four of five do se groups (i.e., 0.8, 4, 20, and500 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 an imals 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 fo r 14 days showed significantly reduced total and free

T 4 levels at ≥ 18 mg/kg/day (Fowles et al. 1994). Serum T 3 and TSH levels and thyroid histology were not evaluated in these single-dose and 14-day mouse studies. Serum levels of total and free T 4 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 Brom kal 70-5 DE, or the congener 2,2',4,4'-tetraBDE (BDE 47), by daily gavage for 14 days (Darner ud and Sinjari 1996; Hallgren and Darnerud 1998; Hallgren et al. 1998, 2001). Exposure to the mixt ures or congener did not cause decreases in serum T 4 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 T 4 were associated with reduced ex vivo binding of T 4 to the plasma thyroid hormone transporter protein transthyretin, as well as with induction of hepatic microsomal enzymes (EROD, MROD, PR OD, 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 corticoster one 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 abst ract (Kuriyama and Chahoud 2003), serum levels of

testosterone and LH were unchanged in adult male offspring of Wist ar 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. Sy mptoms of skin disorders (rashes, acne, increased sensitivity to the sun, darkening or thickening of th e skin, discoloration or deformity of fingernails or toenails, slow healing of cuts) we re 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 sk in disorders (Anderson et al. 1978c). In a medicalPBBs and PBDEs 191

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history survey study conducted in 1976, symptoms of skin disorders (peeling and scaling, erythema, hair loss, increased nail growth, increased sweating) experien ced during the previous 3 years were reported at higher prevalence rates in a group of 321 Michigan r esidents 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 derm al changes in certain species, particularly monkeys, but generally not in haired rode nts. 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 a lterations 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 kera tinization 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 monke ys 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 s quamous metaplasia of sebaceous gl ands and keratinization of hair follicles (Allen et al. 1978). Dermatosis on the ventra l 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 nonspeci fic term used to denote any cutane ous 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

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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 (M oorhead 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 re peated exposure. In general, decreases in food and

water intake are not sufficient to account for decrease s in body weight. Effects on body weight can be quite pronounced following inte rmediate- and chronic-duration exposure, constituting a wastingPBBs and PBDEs 193

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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 FireM aster doses of 400 (BP-6) or 500 (FF-1) mg/kg/day did not affect body weight in rats (Beaudoin 1 977; 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, resp ectively (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 wei ght 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 ga in 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 (M cCormack et al. 1978; Sleight et al. 1978), and 30– 50 mg/kg/day FireMaster FF-1 for 4.5–5 weeks (G upta and Moore 1979; Sleight and Sanger 1976). FireMaster FF-1 dosages ≥ 100 mg/kg/day for 4.5 weeks (lethal do ses) 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 week s, however, did not produce adverse effects on mouse body weight (NTP 1992). Guinea pig, mink, and monke y 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 dosag es 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 (Lambrech t 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 le thal because death was impending between days 33PBBs and PBDEs 194

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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 (contai ning 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 pub lished 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 gr oup 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 P BBs (Bekesi et al. 1978, 1979). In 27 of the 45 Michigan subjects, the peripheral blood ly mphocytes 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

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relative to controls. Assays for membrane markers of peripheral blood lymphocytes showed significant reductions in markers in the Michigan subpopul ation 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 si gnificant differences in ser um PBB levels between the two Michigan subsets. No consistent correlati on could be demonstrated between lymphocyte function and PBB plasma concentration.

Reexamination of a group of 40 Michigan farmers 5 y ears after the first examination (Bekesi et al. 1985; Roboz et al. 1985) showed that the number of T-ly mphocytes 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 classifi ed into three groups acco rding 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 w as similar in the two exposed groups (about 20%). The total leukocyte count, percentage of lymphocytes, and percentage of subp opulations 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 fa rm residents with the highest exposure to PBB had significantly elevated levels of IgM, IgA, and IgG rela tive to Wisconsin dairy farm residents (Bekesi et al. 1985). Cluster analysis of several immunological pa rameters performed for hus bands 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 hi gher 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 dia gnostic findings among the groups.PBBs and PBDEs 196

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The immunological effects of the commercial PBB mixt ures 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 (e xceptions 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. Iden tification of the most sensitive species is further complicated by the fact that not all studies examin ed the same end points, although limited data suggest that guinea pigs may be particularly sensitive. Immuno logical 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 sp leen 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 repor ted 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 (S RBC) (Fraker 1980; Fraker and Aust 1978). This treatment also reduced absolute thymus wei ght by 88% and caused high lethality in mice.

Numerous intermediate-duration studi es have examined the immunological effects of PBBs in rats. For example, treatment of rats with FireMaster FF-1 fo r 25 weeks significantly incr eased absolute and relative spleen weight at ≥ 1 mg/kg/day and significantly decreased ab solute and relative thymus weight at

≥ 0.3 mg/kg/day (NTP 1983). Nevertheless, no hist opathological 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 at rophy 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 fo llowing 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 (Luste r 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 decreasedPBBs and PBDEs 197

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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 th at 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 redu ced (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 we re 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 (G upta et al. 1981). Other studies in mice reported increased lethality (p<0.05) after bacterial inoculatio n 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 Fi reMaster 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 followi ng injection of tetanus t oxoid (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 to tal 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 hum oral and cell immunity relative to a group of 54 unexposed animals (Kateley et al. 1982). The c oncentration of PBBs in tissues from these two cowsPBBs and PBDEs 198

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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 fo r 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 atr ophy, 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 plo tted in Figure 5-2.

Polybrominated Diphenyl Ethers. No studies were located regarding immunological effects in humans after oral exposure to PBDEs. No effects on poke weed mitogen-stimulated DNA proliferation or IgG immunoglobulin synthesis were found in human lymphoc ytes 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 acu te-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 respon se to SRBC (63% of control value, p<0.02) and decreased thymus weight at 72 mg/kg/day (Fowles et al. 1994). Ther e were no exposure-related effects of the 14-day

exposure to ≤ 72 mg/kg/day on natural killer cell (NKC) ac tivity 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 Thuva nder 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 numbe rs of total splenocytes as well as CD4+, CD8+, and CD45R+ cells in spleen (Darnerud and Thuvander 19 98). This was the only dose level of a single congener used in the mice. Rats were not evaluated.PBBs and PBDEs 199

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Histopathological examinations of spleen, thymus , lymph node and/or bone marrow tissues showed no effects of repeated dietary administration in in termediate-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 plo tted 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 c ontamination episode, but positive associations between serum PBB levels and frequenc y of neurological symptoms were not found in several studies. In an epidemiological study conducted by the U.S. Cent er for Disease Control and the Michigan Department of Public Health, fatigue was reported more frequen tly by several putatively exposed groups including 2,148 residents of farms quarantined for PBB contamin ation (36.4% prevalence rate), 1,421 recipients of food from contaminated farms (32.4%), 252 chemi cal workers involved in PBB manufacturing or distribution (22.0%), and 331 residents of farms with lo w levels of PBB contamination (41.4%), than by a small (60 persons) unexposed cont rol group (15.8%); however, no pos itive 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 frequenc ies 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 pos itively 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 re ported 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

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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 con centrating, 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 hi gher 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 be tween PBB oral exposure and the occurrence of neurological effects.

Limited data indicate that orally (gavage) administ ered 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, gr ip 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 observe d 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 a nd/or spinal nerve tissue found no FireMaster FF-1-

related alterations in rats or mice administer ed 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 ass essed in offspring of mice that we re treated with 3 or 10 mg/kg/day doses of FireMaster (FF-1) in corn oil by gavage on every other day during gesta tion and until weaning of the offspring at 21 days of age (Tilson 1992). Acousti c 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 remainingPBBs and PBDEs 201

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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 da ys 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. Neuroc hemical measurements included serot onin and metabolites, dopamine and metabolites, and norepinephrine in the cortex, hippoc ampus, 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 eval uated in offspring of ra ts 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 a nd 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 plo tted 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 neurol ogical 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 in clude functional observational, motor activity, and neuropathology evaluations.

There were no indications of neurotoxicity for comme rcial 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 neurot oxicity in rats and mice exposed to decaBDE in estimated dietary dosesPBBs and PBDEs 202

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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 oppor tunities 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 commercia l 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 o ffspring showed no alterations in motor or sensory development as assessed by overall levels of motor ac tivity, within-session habitu ation of activity, or auditory startle response, although tests of fear c onditioning revealed decreases in cue-based (but not context-based) performance. The lack of motor ac tivity changes in the DE-71 mixture-exposed rats differs from attenuating effects on habituation of mo tor 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 detaile d below, these studies collectively indicate that the developing nervous system is a target of par ticular congeners, as shown by mild performance impairments in tests of spontaneous motor behavi or and learning and memory in adult mice. The aberrations in spontaneous behavior generally appeared to persist and wo rsen with age, and were induced during a defined critical phase of neonatal brain development.

In a series of congener studies using similar experi mental 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; Vi berg 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 lo comotion (horizontal movements), rearing (vertical movements), and total activity (all types of vibratio ns within a test cage). The measurements werePBBs and PBDEs 203

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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 lo cate the platform over a total of 20 trials in a 4-day acquisition period was used as a measure of learning ab ility. 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 (B DE 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-habi tuating 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 animal s, indicating that the effects worsened with age. The findings are consistent with tissue distribution da ta in this study, suggesting that brain uptake of

14 C-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). Additiona l 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 be havior 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, a lthough 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 r esults 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 mazePBBs and PBDEs 204

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performance was tested at 5 months of age in th e 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 thir d period, exposure to 10.5 mg/kg BDE 47 and ≥ 0.8 mg/kg BDE 99 caused significantly more activity than contro ls at 2 and 4 months as shown by dose-related increases in the test measures. The investigators not ed that this nonhabituatin g 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). Th e 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 d esigned 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 th at 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 di minished novelty of the test chambers) was seen in the mice treated at either 3 or 10 days of age. Th ere 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 obser ved for BDE 99 and other congeners in the various

studies with NMRI male mice.PBBs and PBDEs 205

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In a study of perinatal exposure to 2,2’,4,4’,5-pent aBDE (BDE 99), groups of two male and two female CD-1 Swiss female mice were gavage d 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 (bod y weight gain, hair growth, day of eyelid and ear opening, day of incisor eruption) and neurobehavi oral development (righting reflex, forelimb stick grasping reflex, forelimb placing reflexes, negative ge otaxis, 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 duri ng 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 testin g showed changes suggestive of an age-dependent alteration in activity at ≥ 6 mg/kg/day; effects included hyperactiv ity (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 abnorma lities 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 resp onse 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 valu es for PBB levels were not reported.

No relationship was found between ser um levels of PBBs and frequency and duration of lactation in a retrospective study of women exposed to PBBs durin g the Michigan contamination episode (Thomas et al. 2001). A group of 1,020 women with available initial serum PBB levels was identified from theMichigan Department of Community Health PBB registry. Among these participants, 446 had a liveborn infant after their initial serum PBB level; char acteristics 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 deliver y of 5.5±5.2 ppb, duration of breast-feeding as mainPBBs and PBDEs 206

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source of nutrition of 2.6±3.3 months, and total du ration of breast feeding of 4.1±5.3 months. The

numbers of women who breast fed thei r 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 90 th 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 th ree outcomes was significantly associated with serum PBB levels, even after controlling for maternal age, pr evious history of breast-feeding, body mass index, maternal education, household income, history of smoki ng in the year before pregnancy, consumption of alcohol during the first trimester of pregnancy, histor y of thyroid disorder, gestational age of the infant, time to pregnancy, and year of birth.

Effects on reproductive organs and reproductive functi on have been observed in animals following oral exposure to PBBs. An increased incidence of uterin e 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 F 0 rats were fed ≥ 5 mg/kg/day FireMaster BP-6 in the diet from GD 8 through postpartum day 28 (weaning), reproductiv e performance with respect to length of gestation

or litter size was not affected in the F 1 or F 2 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, respec tively, on alternate days between GDs 0 and 14 (Beaudoin 1979). Histological examination of reproduc tive organs in male and female rats and mice revealed no abnormalities following gavage treatment w ith 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 epithelia l 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 provi ding 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 endomet rial 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). Follo wing 6–7 months of exposure to 0.012 mg/kg/dayPBBs and PBDEs 207

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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; prio r 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 require d one to three breedings), but two displayed prolonged implantation bleeding and another two had fetuses th at were aborted or s tillborn (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 plo tted 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 fu nction is limited to negative findings in a onegeneration study of decaBDE (FR-300 BA) in rats (D ow 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 fema le rats were exposed to 3, 30, or 100 mg/kg/day doses in the diet for 60 da ys before mating and subsequently during a 15-day period during which they were mated. Both sexes con tinued to receive the test diet throughout gestation and until the end of the 21-day lactation period. Parameters monitored incl uded length of time between first day of cohabitation and partur ition, numbers of live and dead ne wborn, 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), sk eletal examinations (weanlings), and cytogenetic evaluation of bone marrow (adults and weanlings) we re 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 reporte d 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 puber ty 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 expo sed 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 plo tted 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 p eak 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 de velopment. 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 child ren than control children had frequent upper respiratory illnesses such as colds, r unny noses, and sore throats (Weil et al. 1981).

PBBs were measured in the fat of the infants and in th e 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. Mate rnal blood levels ranged from 0.001 to 0.845 ppm and seven mothers had levels that were not detectable (<0.001 ppm). Se agull (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 fi ve tests between PBB levels in fat tissue and developmental abilities. Mean fat concentrations of PBBs were 0.50 ppm (ra nge, 0.10–0.74 ppm) and 4.218 ppm (range, 0.116–20.960 ppm) in the low and hi gh exposure groups of this study. Schwartz and Rae (1983) later administered the full battery of ne uropsychological 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 negativePBBs and PBDEs 209

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correlations between PBB levels in adipose tissu e (measured in the earlier study) and performance on some of the developmental tasks, but the tasks w ith significant correlations were not the same as those noted in the earlier study by Seagull (1983). The avai lable studies, primarily due to the small data set and the inconsistency of the results, do not conclusive ly establish or eliminate the possibility that in utero and early infancy exposure to PBBs might adversely a ffect the development of human children. The information suggests that if the Michigan PBB c ontamination episode produced any effects on child development, they were subtle.

A comparison of 1966–1981 fetal mortality rates fo r Michigan counties with a high percentage of quarantined farms (6.8–20.4%) with those of Mich igan counties with no quarantined farms did not conclusively establish differences in rates or tre nds after the 1973 contamina tion 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 oc curring after 20 weeks of gestation (early trimester abortions were not counted), and the two populati ons 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, structur al abnormalities, growth retardation, liver effects, and neurological effects in offspring. Developm ental toxicity has been observed in studies with hexabromobiphenyl and octabromobiphenyl comme rcial 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 FireM aster 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 ra ts on one of several days during pregnancy caused increased fetal resorptions, and 400 or 800 mg/kg prod uced 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 liv er 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 tr eated diets as the damsPBBs and PBDEs 210

5. HEALTH EFFECTS

included increased hepatic ALA synthe tase activity (the rate-limiting enzy me in porphyrin synthesis) and increased urinary excretion of uro- and coproporphyr ins 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/da y FireMaster FF-1 for 60 days before breeding until 8 weeks postpartum (4 weeks postweaning) and obser ved for up to the following 2 years included vacuolization and altered foci in the liver (Chhabr a et al. 1993; NTP 1992). Pups of mice that were similarly perinatally exposed to 1.5 mg/kg/day Fire Master 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, an d increased liver weight associated with hepatocyte swelling, vacuolization, and focal necrosis were observed in F 1 generation rats whose only exposure was from the mothers fed a diet providing 5 mg/kg/da y FireMaster FF-1 from day 8 of pregnancy until weaning at 28 days postpartum; less severe liver responses were observed in the F 1 offspring of dams treated with 0.5 mg/kg/day (McCormack et al. 1981). Although survival of F 2 and F 3 generations was not affected by the 5 mg/kg/day treatment of the F 0 rat dams, F 2 offspring, but not F 3 offspring, displayed increased liver weights, liver enzyme induction, an d hepatic histological alterations compared with controls (McCormack et al. 1981). Dietary administra tion of 2.5 mg/kg/day FireMaster BP-6 to rats from gestation day 0 through postpartum day 15 produced in creased relative liver we ights, decreased body weights, and decreased serum levels of the thyroid hormone, T 4 , 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 hepa tic cytochrome P-450 and associated enzymic activities in both dams and pups; a diet providing 0.05 mg/kg/da y produced no hepatic effects in dams, but inducedPBBs 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 sensitiv e than their dams to liver enzyme induction, or that due to the different kinetic parameters am ong 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 biocon centration of PBBs in breast milk (Dent 1978). Performance deficits in tests of ope rant 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-ol d 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 fro m 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 ga in 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 we ight and fetal abnormalities including exencephaly, cleft palate, and hydronephrosis; 5 mg/kg/day did not produce significan t developmental effects in this study (Corbett et al. 1975). Early postnatal deaths occurred among offspr ing 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 th e thyroid, and decreased serum levels of thyroid T 3 and

T 4 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 l actation; 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 paramete rs of immune function in 4-w eek-old offspring of sows fedPBBs and PBDEs 212

5. HEALTH EFFECTS

≤ 2.5 mg/kg/day FireMaster BP-6 during gestation a nd 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 Fire Master 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 am ong first- and second-genera tion 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 signi ficant (p=0.08). Stillbirths and preweaning deaths were not significantly increased, but all mortality was attr ibutable to dystocia. Incidences of dystocia and calf mortality appeared to be related to higher bi rth 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 we re similarly treated with a maternolethal dosage (67 mg/kg/day) of FireMaster BP-6, three aborte d 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 plo tted 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 Laborat ories 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 commerc ial BDE mixtures have been shown to be overtly teratogenic in animals, although neurobehavioral t ests, 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 materi al was a composite of current commercial decaBDE mixtures produced by three manufacturers and ha d a composition of 97.34% decaBDE and 2.66% nona-PBBs and PBDEs 213

5. HEALTH EFFECTS

and octaBDE congeners. The mixture was administer ed to groups of 25 mated female Sprague-Dawley rats by gavage in corn oil in daily doses of 0, 1 00, 300, or 1,000 mg/kg/da y 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, grav id uterine and liver weights, gross lesions, and uterine implantation indices (e.g., numbers of corpora lut ea, 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% d ecaDBE, 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 cont rol values. As discussed in Section 5.2.2.5 (Reproductive Effects), a one-generation study of th e 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 hist ology of pups at weaning (Dow Chemical Co. 1975; Norris et al. 1975b).

Developmental toxicity testing of another octaBDE mixt ure (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 ef fects were observed at ≤ 15 mg/kg/day. The only statistically significant (p<0.05) finding at 25 mg/kg/day was an incr eased serum bromide level. Effects observed at 50 mg/kg/day included significantly reduced mean ma ternal body weight gain during the posttreatmentPBBs 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 c ontrol group but exceeded historical control range), 39% reduced mean fetal weight (p<0.01), skeletal va riations (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 co mmercial 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 statisti cally 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-rela ted 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 postdosag e 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 20 0 mg/kg/day, respectively. There were no exposure-related indications of deve lopmental 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 of fspring 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 T 3 and T 4 hormone

levels showed that serum T 4 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 pr egnancy and birth indices in mice that were gavagedPBBs 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 ne urodevelopmental 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 ra ts that were orally exposed during gestation and lactation (GD 6 to PND 21). In one study, the matern al 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 ey e opening. There were no alterations in any of these indices. The second study also found no changes in mo tor 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 T 4 levels in the offspring and dams (s ee Section 5.2.2.2 Endocrine Effects).

Delays in reproductive development occurred in male a nd female Wistar rats that were exposed to the pentaBDE commercial mixture DE-71 in studies that ar e 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 v esical and ventral prostate weights were reduced at 60 mg/kg/day, although there were no significant cha nges 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 st atistically 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 inform ation 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 chang es in serum levels of thyroid hormones (decreased

T 3 and T 4, increased TSH). The onset of puberty was also delayed in female offspring (vaginal opening), but not in male offspring (preputial separation), of Lo ng 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 a nd 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 a nd Chahoud 2003; Talsness et al. 200 3). 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 leve ls of testosterone and LH (Kuriyama and Chahoud 2003). Effects in the ma le 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), a nd 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 degene rative changes that included vacuolization and accumulation of vesicules in the cytoplasm of the stroma l 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 plo tted in Figure 5-3.

5.2.2.7 Cancer

Polybrominated Biphenyls. No epidemiological studies were lo cated that provided evidence for an association between exposure to PBBs and the occu rrence 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 th e EPA established a cohort of Michigan residents with varying levels of PBB exposur e to determine the short- and long- term effects (especially cancer) of exposure to PBBs (Landrigan et al. 1979). The epidem iological and clinical data collected during the first 4 years after the Michigan PBB contamination ep isode 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 co hort was divided into seven groups based on serumPBBs and PBDEs 217

5. HEALTH EFFECTS

PBB levels, no trend with concentra tion 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 (C EA), 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 diffe rence was not statistically significant (Anderson et al. 1978b). Cancer was not listed as a condition in th e report of results of a symptomatology survey completed by this cohort (Anderson et al. 1979). Repo rts of follow-up examinations of this cohort have not been reported.

A relationship between serum PBBs and risk of br east cancer was suggested in a nested case-control study of 1,925 women enrolled in the Michigan Departme nt of Public Health registry for persons exposed to PBBs (Henderson et al. 1995). Study particip ants had lived on or received food from a farm quarantined by the Michigan Department of Agri culture, were recruited from July 1976 to December 1977, and followed up annually fro m 1978 through 1993. Twenty wome n who developed breast cancer were age- and race-matched to 290 controls. Medi an 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). Cond itional 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 slight ly 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 va riable was used in the analysis (OR=3.3, 95% CI=0.9-11.4). The results were essen tially the same when the data were ad justed 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 b ased 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 interv als 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., exposu re to other organochlorin e chemicals and estrogen receptor status).PBBs and PBDEs 218

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Another study of the Michigan PBB registry evalua ted associations between levels of serum PBBs and risks of various site-specific can cers (Hoque et al. 1998). Primary cancers (195 malignancies) were identified in 187 persons among 3,899 registrants enrolle d 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 me asured 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 90 th and 95 th percentiles. Conditional logistic regression was used to calculate univariate and multivariate (adjuste d) 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 sp ecified) (8 cases) showed increasing dose-response relationships for risk as PBB concentrations incr eased. 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 hi story 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 cate gory (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 fo r 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 hea lth 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 ti ssue 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 ne oplastic 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 we re examined at 2, 6, 10, and 14 months following treatment (Kimbrough et al. 1978b). Neoplastic nodules we re 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 we re 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 Sh erman 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 carcinom as and liver neoplastic nodules in offspring that were observed for 2 year s (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 inciden ce of nodules was 85.7% (24/28 versus 1/25). In repeated dose studies performed by NTP (1983), Fire Master 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 neopl astic 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 hepatocellu lar 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 hepatobl astomas were observed. Thyroid follicular cell adenoma tended to increase in treated female mice, but data are inconc lusive due to low incidences and small numbers of

animals.

The carcinogenicity of FireMaster FF-1 was additiona lly 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 exposu re involved dietary administration of PBBs (F 1 diets)

to ≈ 8-week-old animals for up to 104 weeks (rats) or 105 weeks (mice). Perinatal-only exposure involved dietary treatment of dams (F 0 diets) for 60 days prior to breeding and throughout gestation and lactation until pups were 8 weeks old. The pups were admini stered the same treatment as the dams from weaning at week 4 until age 8 weeks, and were subsequently ad ministered the same or different dietary treatments (F 1 diets) for up to 104 weeks (rats) or 105 weeks (m ice). 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 peri natal and adult exposure regimen.

Eight F 0 :F 1 doses (estimated) were tested in rats amo ng 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 peri natal-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 tumo rs 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 carcinomaPBBs 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 car cinoma (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 perina tal 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 oc curred 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 ex posed 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 do ses. 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 incide nces 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 experime ntal groups indicates that significant increases in the incidence of the leukemia are associated with increasing F 1 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 F 0

exposure.

Eight F 0 :F 1 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 (c ombined) were 5 of 50, 42 of 50, and 47 of 48, respectively. Combined perinatal and adult exposur e resulted in increased inci dences 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 possi ble enhancing effect of combined perinatal and adult exposure could not be adequately assessed in ei ther 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 (combine d) in females (50 of 50, p<0.001). Compared to 0:3.9 mg/kg/day adult-only exposure, combined peri natal 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 expo sure also caused significan tly 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, carcinom a, 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-on ly 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 nonhe patic tumors in mice of either sex.

The existing evidence conclusively demonstrates that th e 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 carcinom a/adenoma ratio), and that mice are more sensitive than rats (based on earlier occurrence of hepatocellula r 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 he patocellular 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, devel opment of enzyme-altered hepatic foci (putative preneoplastic lesions) was assessed in rats that were 70% hepatectomized, initiated with a subcarcinogenic intraperitoneal dose of diethylnitro samine 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 initia tion with dimethylnitrosamine (NDMA) or N-nitrosopyrrolidine (NPYR) promoted development of hepa tic 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 gr oup of hamsters given a single subc utaneous 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 tu mor promoting activity in rats that were partially hepatectomized and initiated with diethylnitrosamin e (DENA). Numbers of hepatic GGT-altered foci and/or neoplastic nodules were increased following pr omotion 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 developmen t 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 in formation on structure-promotion relationships for

PBBs is discussed in Section 5.5.2.

The tumor initiating potential of PBBs is not well ch aracterized. Numbers of GGT-altered foci were significantly increased in partially hepatectomized ra ts 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 ac tivity in hepatocarcinogenesis. The potential forPBBs and PBDEs 224

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liver tumor initiation by PBBs appears to be w eak 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'-tetraBD E 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 w ith 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 regr ession, adjusted for age, gender, sum of PCBs, and sum of chlordanes, was performed on cases and contro ls in three concentra tion 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 th e carcinogenicity of decaBDE in animals. No carcinogenicity studies of octaBDE or pentaBDE were located in the available literature.

The National Toxicology Program evaluated the car cinogenicity 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 examin ations were performed on all animals in all dose groups including those that were moribund or died dur ing the study. Reported estimated dose levels in the rats were 1,120 and 2,240 mg/kg/day in males a nd 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 tr ends (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 c ontrols. The increased incidences of neoplastic nodules were considered as “some evidence of carcinogeni city” in both sexes. However, although it was concluded that there was some evidence of carcinogenic ity in male and female rats based on “neoplastic nodules”, this is a poorly defined a nd understood term that is no longer used by NTP to characterize hepatoproliferative lesions in rats. A dose-relate d trend for mononuclear cell leukemia was observed in

treated male rats but was not considered to be bi ologically 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). Hepatocellula r adenoma or carcinoma (combined) occurred at significantly increased incidences in low-dose ma le 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 pos itive 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 summari zed 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 histol ogical examinations showed no exposure-related neoplastic effects. The ability of th is 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 th e 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 wo rkers 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 appe ar to have had a high pot ential 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 im portant 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 am ong a group of four rabbits exposed to up to 10,000 mg/kg of body weight of a commercial octabr omobiphenyl 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 hexabr omobiphenyl mixture in the same vehicle as the octabromobiphenyl mixture. A dose of 10,000 mg/kg a pplied for 24 hours killed two of four rabbits. The cause of death was not reported. A commercial mixtur e 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 am ong the hexa-, octa-, and decabromobiphenyl mixtures reflect differences in le thal 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.LOAELTable 5-5 Levels of Significant Exposure to Polybrominated Biphenyls - DermalSpecies(Strain)SystemReferenceNOAEL Less Serious Serious Chemical FormExposure/Duration/Frequency(Specific Route)ACUTE EXPOSUREDeath(NZW)Rabbit10000mg/kg/daymg/kg/dayWaritz et al. 1977(OBB)212024 hr(NZW)Rabbit(4/4 died)5000mg/kg/daymg/kg/dayWaritz et al. 1977(HBB)23324 hrSystemic(Albino)RabbitDermal658mg/kg/daymg/kg/dayMillischer et al. 1980(DBB)389once(NZW)RabbitDermal(hyperkeratosis)0.19mg/kg/daymg/kg/dayNeedham et al. 1982(FF-1)2725 d1x/d(NZW)RabbitHepaticM100mg/kg/daymg/kg/day(increased liver weight;necrotic foci)1000mg/kg/daymg/kg/dayWaritz et al. 1977(HBB)23424 hrBd WtM1000mg/kg/daymg/kg/day(11% weight loss)5000mg/kg/daymg/kg/day(NS)RabbitHepatic(increased liver weight)1mg/kg/daymg/kg/dayWaritz et al. 1977(OBB)2402 wk5 d/wk1x/dDermal1mg/kg/daymg/kg/dayBd Wt1mg/kg/daymg/kg/dayLOAELTable 5-5 Levels of Significant Exposure to Polybrominated Biphenyls - Dermal(continued)Species(Strain)SystemReferenceNOAEL Less Serious Serious Chemical FormExposure/Duration/Frequency(Specific Route)INTERMEDIATE EXPOSURESystemic(AlbinoHartley)Gn PigDermalM62mg/kg/daymg/kg/dayBd Wt = body weight; d = days; DBB = deca-brominated biphenyl; Derm = dermal; FF-1 = FireMaster FF-1; HBB = hexa-brominated biphenyl; hr = hour(s); M = male; OBB =octa-brominated biphenyl; wk = week(s); x = time(s)Waritz et al. 1977(OBB)2383 wk3x/wkPBBs 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 human s 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 an d 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 si ngle 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 pe rformed. 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 significan t increase (p<0.01) in relative liver weight was reported in rabbits after application of 1 mg/kg/da y of a commercial mixture of octabromobiphenyl in

corn oil to the intact and occluded shaved dorsal sk in 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

DecaBDE 2101 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 hrPBBs and PBDEs 231

5. HEALTH EFFECTS

No studies were located regarding hepatic effects fo llowing 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 onl y chemicals manufactured and who had volunteered for a comprehensive medical evaluation performed 3 m onths 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 r esults of this study suggest that occupational exposure to PBBs, decaBDE, and/or bromine affected the thyroi d, but the mixed chemical exposure and a lack of data on serum or tissue levels of the chemicals precl ude attributing effects solely to any particular congener or mixture of congeners.

No studies were located regarding endocrine effect s 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 effect s in animals after dermal exposure to PBDEs.

Dermal Effects.

Polybrominated Biphenyls. As discussed in Section 5.2.1.2, resu lts from a medical history survey study of workers in a PBB manufacturing plant and a nonexpo sed 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 e xposure, 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 mi xture 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 vehi cle, 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 eviden ce 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 aff ect 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 (am ount not reported) to occluded intact or abraded skin caused no skin response, but the water-moistene d formulation caused slight and transient erythema (Norris et al. 1975a). None of these studies reported th e 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 octabromobi phenyl 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 equi valent 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 commerc ial formulation of octabromobiphenyl did not induce bromacne when applied to the ear of rabbits for 30 da ys (Norris et al. 1975a). Only slight erythema and exfoliation was observed. Doses of 62 mg/kg of oc tabromobiphenyl 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 (N orris 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 re ported 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 so lid 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 ra bbit skin (IRCD 1975a, 1975b; Norris et al. 1975a). A similar single application of solid 77.4% decaBDE (2 1.8% nonaBDE and 0.8% octa BDE) 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-sen sitizing 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 adjuva nt and corn oil, (2) 2.5% octaBDE or 5% pentaBDE solutions in corn oil, and (3) 2. 5% octaBDE or 5% pentaBDE in th e 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 we re 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 res ponses in any of the animals, indicating that the PBDEs did not cause delayed contact hypersensitivity.

A 10% chloroform solution of 77.4% decaBDE (cont aining 21.8% nonaBDE and 0. 8% octaBDE) did not induce bromacne when applied to the ear of rabb its for 30 days (Norris et al. 1975b). A slight erythematous response and slight exfoliation were th e 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 membran es was observed after a single instillation of an unreported amount of dry solid octabr omobiphenyl mixture to the eye in rabbits, but the cornea, iris, and lens were unaffected (Norris et al. 1975a). Co mmercial grade decabromobiphenyl did not cause eye irritation in rabbits when 0.05 mg in olive oil w as instilled for 30 seconds fo llowed by rinsing, but application of an unspecified amount of dry powder w ithout rinsing was slightly irritating (Millischer et al. 1980). Mild conjunctival redness and swelling and a copious discharge was reported after applicationPBBs and PBDEs 234

5. HEALTH EFFECTS

of 100 mg of an unspecified commercial PBB powde r mixture (either hexa- or octabromobiphenyl) for 20 seconds into the conjunctival sac of two rabbits (War itz et al. 1977). These effects disappeared within 4 hours in both washed (with tap water) and unwash ed 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), 1 00 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 re ported 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 comme rcial 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 do se 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 rabb its 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, covere d 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 parame ters 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 ou t. The subjects had worked directly with PBBs during the previous 5 years. Immunological studi es 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 unexpo sed group. One of three blastogenic responses (PWM) was significantly reduced (p<0.01) relative to c ontrols, but was within the normal range for the laboratory. No specific information was provided re garding the skin testing and immunoglobulin levels.

No studies were located regarding immunological e ffects 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 eff ects 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 manufacturi ng 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 in dividual serum PBB values were not reported.

No studies were located regarding reproductive eff ects 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 applica tion 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 stud y 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 atPBBs 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 c ontrols. Toxic effects included mortality, which caused the dose reduction after 5 weeks, and sev ere 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 sugg ests that PBBs acted as a promoter rather than directly causing DNA damage (initiator). A commerc ial 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 tr eated (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 test ed 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; Ka vanagh 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 CytogenicityRat bone marrow cells (onegeneration reproduction study) Chromosome aberration – Norris et al. 1973, 1975 (77.4% decaBDE, 21.8% nonaBDE)

– = negative result; DBB = decabromobiphenyl commercia l mixture; decaBDE = decabromodiphenyl ether; DNA = deoxyribonucleic acid; FF-1 = FireMast er FF-1; nonaBDE = nonabromodiphenyl ether; PBBs = polybrominated biph enyls; PBDEs = polybromi nated diphenyl ethersPBBs 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'-tetrabr omodiphenyl 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 = hexabr omobiphenyl (unspecified); PBB = unspecified mixture; PBBs = polybrominated biphenyls; PBDEs = polybrominatedd diphenyl ethersPBBs and PBDEs 241

5. HEALTH EFFECTS

An in vitro study with a 14 C-PBB mixture containing 12 major components found only traces of radioactivity bound to rat liver microsomal macrom olecules (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 th at no radioactivity was covalently bound to DNA following incubation with 14 C-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 com pounds 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 publis hed 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) fo r 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 in duce 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 Chine se hamster SPD8 and Sp5 V79 cells (Helleday et al. 1999).

5.4 TOXICOKINETICS

Data regarding the toxicokinetics of PBBs in human s are limited to information derived from cases of accidental ingestion of food contaminated with PBBs and cases of occupational exposure by thePBBs and PBDEs 242

5. HEALTH EFFECTS

inhalation and dermal routes. These data provide qua litative evidence that PBBs are absorbed in humans by the inhalation, oral, and dermal routes. Limite d quantitative data in animals indicate that some PBB congeners are well absorbed after oral exposure. De rmal absorption data for an imals are insufficient for estimating absorption rates, and no inhalation absorpti on 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 an imal 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 th e 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 metabolis m of some PBB congeners depends on the bromine substitution pattern. PBB conge ners of low bromine content are transformed into hydroxylated derivatives that are predominately eliminat ed 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 ab sorption 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 resi dues 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 absorpti on of PBBs in animals after inhalation exposure to PBBs. However, increased bromine concentrations we re 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 regardi ng absorption of PBDEs in humans after inhalation exposure. Evidence for the inhalatio n 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 LakesPBBs and PBDEs 243

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Chemical Corporation 2001a, 2001b). The absorptio n 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/m 3 for 6 hours/day, 5 days/week, for 13 weeks.

5.4.1.2 Oral Exposure

Polybrominated Biphenyls. Quantitative oral absorption data in hum ans 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 P BBs 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 quantita tive data exist. By comparing the amount of radioactivity in the feces of rats administered a single 1 mg/kg oral dose of 14 C-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 la ter 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 hexabrom obiphenyl congener, a comme rcial 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 indica tes 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 regard ing 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 lowe r brominated PBDEs ar e readily absorbed .

Studies with 14 C-labeled decaBDE (BDE 209) in rats consistent ly indicate that gastrointestinal absorption of this congener is low at 10% or less in rats (E l 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 aPBBs and PBDEs 244

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single 1 mg/kg dose of 14 C-decaBDE, administered as a former lo w purity commercial mixture (77.4% decaBDE, 21.8% nonaBDE, 0.8% octaBD E) 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 hi gher purity commercial mixture as unlabeled decaBDE (92% pure) on days 1–7 and 14 C-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 14 C-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 radi oactivity in the feces ranged from 82.5±4.7%

to 86.4±8.5% of the dose and was not related to dose le vel or time of sacrifice (24, 48, or 72 hours after

14 C-decaBDE intake). For both dose levels, the percent of 14 C 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 ti ssue concentrations following intravenous and oral administration, NTP (1986) estimated that oral abso rption was 0.33±0.19% at the highest dietary level (50,000 ppm).

A study of laboratory synthesized 14 C-decaBDE (>98% pure) was conducted in which a single 3 μmol/kg ( ≈ 3 mg/kg) dose was administered to normal or b ile duct-cannulated male Sprague-Dawley rats by gavage (Klasson Wehler et al. 2001; Morck a nd 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 absorpti on. 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 investig ators 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 dietsPBBs and PBDEs 245

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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-pentaB DE (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 neglig ible, 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 followi ng eight congeners: 2,2’,4,4’,5,5’-hexaBDE (BDE153), 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 14 C-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 c ongener in the feces in 24 hours. The investigators concluded that these values represented the non-abso rbed doses, indicating that absorption was 93–95%.

5.4.1.3 Dermal Exposure

Polybrominated Biphenyls. No studies were located regarding abso rption 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 i nhalation and/or oral expos ure cannot be ruled out.

Similarly, dermal absorption in rabbits can be inferre d from reports of lethality and liver effects observed after application of a commercial mixture of hexa bromobiphenyl to abraded and occluded dorsal skin (Waritz et al. 1977).PBBs and PBDEs 246

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Polybrominated Diphenyl Ethers. No information was located regarding dermal absorption of PBDEs in humans. The only information regarding dermal absorp tion in animals is that from a study of absorption

in an in vitro preparation (Hughes et al. 2001). In that study, 14 C-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 radioactiv ity. Two to 20% of the radioactivity was found in the skin, and the lowest dose applied had the highest pe rcentage of the dose in th e 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 distributi on 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 octabromobi phenyl 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 a nd 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/m 3 of dusts of octaBDE 8 hours/day for 14 days. At necrops y, sections of the lungs, adipose tissue, and liver were collected for bromine analysis using a ne utron 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 concentr ation-related. In the liver, the concentration of bromine was also elevated in all groups relative to controls except in the 1.2 mg/m 3 exposure group; the

increases in the liver were not as marked as in the lungs or in adipose tissue.PBBs and PBDEs 247

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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 correla tions were determined (Eyster et al. 1983). For example, in parturient women from Michigan, statis tically 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 add ition, there was a significant correlation between PBB

levels in serum and feces and between serum and bilia ry fluid samples in farmers and chemical workers in Michigan. In groups of pregnant, nonpregnant, an d male chemical workers th e 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 consiste nt with other reports regarding Michigan populations (Landrigan et al. 1979; Wolff et al. 1982). It is unc lear why the partitioning ratios between male chemical workers and farmers should differ. The investigators not ed that the group of farmers was much larger and might represent a better sample, as well as the possi bilities that the farmers may have been more physically active or, for a variety of reasons, may ha ve 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 subj ects in the Grand Rapids, Michigan area indicated that renal fat had the highest single PBB concentration (1.65 μ g/g wet weight) and the highest mean concentration (0.475 μ g/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 vent ricle), and low (ratios of 0.02–0.09, kidney, lung, brain, skeletal muscle, thyroid, and nonatheromatous aorta).

As with the structurally related PCBs (Agency fo r 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

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In rats treated by gavage wi th one or four daily doses of 14 C-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 14 C-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 pr edicted by a physiological compartment model (Tuey and Matthews 1980). When the model was scaled to nonlactating huma ns as discussed in Section 5.4.5 (Physiologically Based Pharmacokinetic/Pharmacodynamic Models), hum an 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 c ohort (Lambert et al. 1990) (see Section 5.8.1).

In rats fed diets containing octabromobiphenyl mixt ure for several weeks, adipose tissue and liver

accumulated much more bromine than did skeletal mu scle (Lee et al. 1975b). For example, after 2 weeks of treatment, adipose of rats dosed with 50 mg/kg/da y had 200 times more bromine than did adipose of control rats; the liver of these rats had 100 times mo re 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 we re 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 tissu e, heart, and skin at levels ranging from 0.14 to 0.25% of the administered dose; the liver, pa ncreas, 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, p eaked 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 afte r 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

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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. Accordi ng 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 Michig an 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 fo r up to 11 months, the concentration of PBB residues in adipose tissue were higher than in brain and sk eletal muscle at all times (Aulerich and Ringer 1979; Ringer et al. 1981). The source of the PBBs (FireMas ter FF-1 versus food contaminated with PBBs) did not seem to have a significant influence on the qua litative 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 PB B 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 ad ipose 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 tissu es 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 c ontrast, PBB levels in liver and body fat of guinea pigs exposed briefly through lactation were cons iderably 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 singlePBBs and PBDEs 250

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gavage dose of 1 mg/kg of a 14 C-labeled former commercial decaBD E mixture (77.4% pure containing 21.8% nonaBDE and 0.8% octaDBE), radioactivity co uld be detected on day 1 in all sampled tissues (adipose, skin, liver, heart, adrenals, spleen, pancr eas) (Norris et al. 1975b). On day 16 after dosing radioactivity was only detected in ad renals 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 th e 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 liv er, the bromine content was elevated on day 0 of the recovery period, but decreased and stabilized n ear 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 incr eased 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 co mpared to controls (Norris et al. 1975b). Placental transfer cannot be totally ruled out si nce 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 14 C-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 14 C was recovered in the feces dose in 72 hours. In the second study, rats were exposed to dietar y concentrations of 277 or 48,000 ppm ( ≈ 20 or 4,300 mg/kg/day), and fecal recovery of 14 C 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 14 C activity detected in the body at any time was only ≈ 1% of the dose. In general, the highest amounts of 14 C 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 14 C in the gastrointestinal tract, followed by liver, kidney, lung, skin, and adipose.PBBs and PBDEs 251

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A study of laboratory synthesized 14 C-decaBDE (>98% pure) was conducted in which a single 3 μmol/kg ( ≈ 3 mg/kg) dose was administered to male Sprague-D awley rats by gavage (Kl asson 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 formul ated to enhance solubility and optimize absorption. The amount of the 14 C dose remaining in the body after 3 and 7 days was approximately 9% (calculated by subtracting to tal urine and feces output from 100%). The highest concentrations of radioactivity on a fresh weight b asis 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 decaB DE is not readily distributed to adipose. The investigators speculated that decaDBD does not parti tion into lipids and is tr ansported thr ough aqueous compartments (e.g., serum and bile) due to binding to transport proteins. As discussed below, PBDEs with fewer numbers of bromine preferentially accu mulate in adipose. The accumulation of lower brominated congeners appears to be due to their par titioning and retention in lipid-rich tissues, as well as

rates of metabolism and elimination th at 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 sing le 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 perirena l fat. Distribution from extra-adipose tissues into the fat was essentially complete after 4 days. Five co ngeners 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. Th e mean half-life of the tetraBDE congener was 19.1 days in males and 29.9 days in females and signi ficantly (p=0.01) different between the sexes. The mean half-lives of the penta- and hexaBDE congene rs 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.PBBs and PBDEs 252

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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 pent aBDE mixture assessed the following six congeners: 2,2’,4,4’-tetraBDE (BDE 47), 2,2’,3,4,4’-penta BDE (BDE 85), 2,2’,4,4’,5-pentaBDE (BDE 99), 2,2’,4,4’,6-pentaBDE (BDE 100), .2,2’,4,4’,5,5’-he xaBDE (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 resemb led that of the commercial pentaBDE mixture, suggesting that there was no preferential tissue accumu lation of the tested congeners. The study of the octaBDE mixture assessed the following eight c ongeners: 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 hept aBDE, and two of the unknown octaBDEs) were not detected in the liver, and one congener (an unknow n 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 in creasing bromination; total retentions in the liver and carcass were 64.4% for one of the hexaBDEs (B DE 153), 27.0–37.1% for the three heptaBDEs, and

16.2–22.8% for two of the octaBD Es. 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 14 C-2,2’,3,3’,4,4’,5,5’,6,6’-decaDBE (BDE 209), or 0.8 mg/kg

of 14 C-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, respectiv ely, whereas the brain of those exposed on PND 19 had only 0.6% of administered amount (Viberg et al. 20 01b, 2003a). Seven days after exposure, the amount of radioactivity in the brain had increased approximat ely 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 14 C-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 w as 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

14 C-BDE 209, the amount of radioactivity in the brain was between 3.7 and 5.1% of the total administeredPBBs and PBDEs 253

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amount in the three different age categories, wh ereas 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 14 C-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 muscl e showed that the majority of the 14 C was in the skin. When deposition was expressed on a concentration basis, the hi ghest 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 14 C-2,2',4,4'-tetraBDE (BDE 47) in corn oil was administered by gavage to rats and mice (Örn and Klasson-Wehler 1998). Approxima tely 86% of the absorbed dose in rats and 80% of the dose in mice remained in tissues (liver, lung, kidne y, brain) at 5 days, mainly in the adipose. In rats, adipose tissue had an approximately 70 times highe r 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 twi ce that in the liver and kidneys. In mice, adipose tissues had an approximately 10-fold higher level of la bel than other tissues on a fresh weight basis. On a lipid weight basis, adipose tissue had a concentration of 14 C 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

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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

14 C-2,2',4,4',5,5'-hexabromobiphenyl (BB-153) adi pose, 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 ad ipose tissue. The percent of the dose remaining in liver and muscle on day 42 was 0.8 and 3.5%, respectiv ely. 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 PB B concentrations at week 6, and these levels were maintained throughout the 36-week observation period. C oncentrations of PBBs were also elevated in the liver, lungs, and pituitary at week 6, whereas PBB leve ls 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 h eart, 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 tissu e. The investigators estimated that, given the elimination half-time from fat of 69 weeks, >1 μ g/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 injecti on 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 th e 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 th e dose. In a different experimental series, thePBBs and PBDEs 255

5. HEALTH EFFECTS

investigators (Bleavins et al. 1980) also showed th at the dam's milk was the major route of offspring exposure since PBB levels, on a per gram or per kit basi s, were significantly higher in 2-week-old kits

than in newborn kits. The ratio of the 2-week P BB 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 Mi chigan 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 lowe r 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 th e groups were also apparent. The differences in peak profiles between farmers and chemical workers we re 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 reducti on (36–52%) in the concentration of PBBs in pressure-cooked meat relative to raw meat (due to loss of fat) has been re ported (Zabik et al. 1978). The decreased heptabromobiphenyl peak in farmers and worker s 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 cont amination episode occurred primarily through consumption of contaminated meat and dairy products. The limited in formation available regarding the metabolism of PBBs in dairy cattle is insufficien t to ascertain whether humans ingested PBBs or metabolic products of PBBs. In a c ontrolled 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, orPBBs and PBDEs 256

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-biphe nylol (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 metabo lism of BB 15 occurs through the formation of an arene oxide. The major urinary metabolite of FireMa ster BP-6 was a pentabromobiphenylol (1% of the dose), which could have resulted from direct hydroxyl ation of the minor pentabromobiphenylol isomers in FireMaster BP-6 or by debromination/hydroxylatio n of the major congener, 2,2',4,4',5,5'-hexabromobiphenyl (BB 153).

In rabbits, metabolism of 2-bromobiphenyl yielded tw o polar metabolites, one metabolite was identifiedas 2'-bromo-4-biphenylol (1% of the dose), and the other metabolite (traces) was also a monohydroxylated derivative, but the position of the hydrox yl 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-bromobiph enyl suggest that the metabolism of this congener

involves the formation of an arene oxide.

Similar results have been reported in rats (Sparli ng et al. 1980). 4'-Bromo-4-biphenylol was the major metabolite of 4-bromobiphenyl (BB 3). 2-Bromobiphe nyl (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) toPBBs and PBDEs 257

5. HEALTH EFFECTS

the substituents; and the 2- and 3- hydroxylated products are subseq uently dehydroxylated, whereas the 4'-hydroxy congener is not.

In contrast to the lower brominated congeners, no ma jor 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 metabo lite 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 he xabromobiphenyl 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) ha ve 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 molecu le 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 tetrab rominated), 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. ItPBBs and PBDEs 258

5. HEALTH EFFECTS

was concluded that the rates of metabolism of P BB congeners depends on the position of the bromines and the form of the cytochrome P-450 induced. Th e ability to metabolize PBBs also depends on the species. For example, hepatic microsomes isolated fro m 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 i ndicating 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 puri ty mixture as unlabeled decaBDE (92% pure) on days

1–7 and 14 C-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 14 C-decaBDE). For both dose levels, the percent of 14 C dose remaining in the gut contents (< 4%) and gut contents (<0.04%) decreased

with time. Of the radioactivit y recovered, >99% was in the f eces and gut contents. HPLC and UV spectral analyses of extracts of feces collected on da ys 9–11 indicated the presence of decaBDE (81% of recovered radioactivity) and three main unidentified metabolites. Th e total radioactivity present as the metabolites tended to increase with increasing dietary le vel of decaBDE (1.5 and 27.9% of total recovery at the low and high dose, respectiv ely). Since absorption of decaBDE was minimal, El Dareer et al. (1987) speculated that the greater extent of metabolis m in the high dose rats may be due to induction of hepatic metabolizing enzymes, or gastrointestinal tr act bacteria, during the period of feeding unlabeled decaBDE. Analysis of feces from rats following a single 1.07 mg/kg intravenous dose of 14 C-decaBDE similarly indicated that the excreted material w as 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 th e oral and intravenous studies. A single metabolite was detected in the bile of rats following a single 0.9 mg/kg intravenous dose of 14 C-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 14 C-decaBDE (>98% pure) was conducted in which a single 3 μ mol/kg ( ≈ 3 mg/kg) dose was administered to normal or b ile 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 solubilityPBBs and PBDEs 259

5. HEALTH EFFECTS

and optimize absorption. Feces, bile, and several tis sues (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 th e normal rats, approximately 90% of the dose was excreted in the feces in 3 days, mainly as metabolit es (65% of dose). In the the two bile duct-cannulated rats, an average of 9.5% of the dose was excreted vi a the bile in 3 days, almost all of which represented

metabolites. Metabolites were characterized as none xtractable, 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 whic h 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. Analys is 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 debromina tion 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 pl ay a role in some of this metabolism. PBDEs with fewer bromines probably have lower rates of metabol ism (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 commercia l 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 formul ation were 2,2’,4,4’-tetraBDE (BDE 47) and 2,2’,4,4’,5-pentaBDE (BDE 99). The octaBDE form ulation was mainly comprised of hexaBDE throughPBBs and PBDEs 260

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nonaBDE congeners and only containe d 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 follo wing six congeners: 2,2’,4,4’-tetraBDE (BDE 47), 2,2’,3,4,4’-pentaBDE (BDE 85), 2,2’,4,4’,5-pent aBDE (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 congener s 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). F our 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. Un like the lower brominated congeners (tetra- to hexaBDEs) in the study of the pentaBDE mixtur e, 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 th at most of the congeners were

metabolized to some extent

The metabolism of 14 C-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 el imination as shown by fecal recovery of 43 and 86%

of the administered 14 C 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, resp ectively). 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 pent aBDEs. The possible thiol-substituted pentaBDEsPBBs and PBDEs 261

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could have been formed through the mercapturic aci d pathway, but no glucuronide or sulfate conjugates

were found in the bile. Evidence for reactive intermed iates in the feces of normal rats was indicated by high nonextractable fractions ranging from 18–52%. Cumu lative excretion of radiolabel in the urine at 72 hours, consisting exclusively of uncharacterized me tabolites, 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; th e 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 14 C-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 14 C ( ≈ 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-tetraB DE. 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 14 C in mice ( ≈ 70% of the parent compound/metabolite fraction) was the metabolite frac tion, which contained the same six metabolites characterized in the rat feces. Approximately 20% of the mouse urinary 14 C 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 me tabolites 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 h ydroxyl-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

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5.4.4.2 Oral Exposure

Polybrominated Biphenyls. No studies were located that provide information on percentage of ingested PBBs excreted by humans. Howeve r, PBBs in biliary fluid of a group of farmers and chemical workers ranged from undetected to 70 μ g/L, and the correlation between serum PBB levels and levels in bile was statistically significant (Eyster et al. 1983). Simila rly, PBB levels in feces ranged from undetected to

862 μ g/kg, and the correlation between serum PBB levels a nd fecal levels was also statistically significant (Eyster et al. 1983).

Serum half-life values have been estimated using hum an 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 leve ls 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 tw o measurements taken when they were not pregnant (Blanck et al. 2000b). The mean initial PBB level wa s 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 tim e. Assuming that PBBs reached equilibrium in the body before substantial amounts we re eliminated and before the firs t 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 incl uding 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 w ith 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 statisticalPBBs and PBDEs 263

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significance (p=0.06). Breast feeding as either a c ontinuous variable or as categorized by tertiles of duration (<3, 3–9, or >9 months), age, and sm oking 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 μ g/kg, with a median of 250 μ g/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 th an in serum. Also, adipose PBB levels are 1.1– 1.5 times higher than the breast milk levels when milk levels were ≥ 100 μ g/kg.

The importance of PBB transfer through lactation in e xperimental animals is disc ussed in Section 5.4.2.2.

There is limited information regarding excretion of P BBs in experimental animals. Rats dosed once with

14 C-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 (M atthews 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 an d persistent (Safe et al. 1978) . In rats gavaged with

14 C-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 th e administered dose was ex creted 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 14 C-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 prop ortion 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 14 C-labeled decaBDE consistently show that this PBDE is rapidly and predominantly eliminated in th e 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 byPBBs and PBDEs 264

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gavage with a single 1 mg/kg dose of a low purity former commercial mixture (77.4% 14 C-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 detect ed in the urine and expired air (Norris et al. 1973, 1975c). No significant differences in excretion pa tterns were observed between males and females. Approximately 91% of the decaBDE-derived radioactiv ity 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 expo sed to a higher purity commercial mixture as unlabeled decaBDE (92% pure) on days 1–7 and 14 C-decaBDE (98.9% pure) on day 8, follo wed 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 14 C-decaBDE intake and was not related to dose level. In the second study, rats we re exposed to dietary concentrations of 277 or 48,000 ppm ( ≈ 25 or 4,300 mg/kg/day). Recovery of radio activity in the feces ranged from 82.5±4.7 to

86.4±8.5% of the dose and was not related to dose leve l or time of sacrifice (24, 48, or 72 hours after

14 C-decaBDE intake). For both dose levels, the percent of 14 C 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 14 C accounted for ≤ 0.01% of the dose.

In a study of laboratory synthesized 14 C-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, approxi mately 90% (86–93%, n=8) of the dose was found in the feces after 3 days; cumulative recovery after 7 da ys was 91% (87-95%, n=4). For the two bile ductcannulated rats, an average of 88 and 9.5% of the do se was recovered in the feces and bile, respectively, within 3 days. There was a large variation in the f ecal excretion of the two cannulated rats that might

have been due to the fact that no bile salts were added to compensat e for the collected bile. Urinary

excretion of 14 C was negligible for all groups at <0.1% of the dose. DecaBDE was speculated to have higher rates of metabolism and elimination than PBDE s with fewer numbers of bromines because it is unlikely to partition into lipids a nd is transported throu gh aqueous compartments (e.g., serum and bile), whereas lower brominated congeners preferentially accu mulate in the body due to their partitioning and retention in lipid-rich tissues.PBBs and PBDEs 265

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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-pentaB DE (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 conge ners ranged from 7.6 to 15.8% of the dose. The study of the octaBDE mixture assessed the following eight congene rs: 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 he xaBDE 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 14 C-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-cannulat ed 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 14 C-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 excret ed 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

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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 14 C-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 ex creted in the urine. Two decay components were calculated from excreti on 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 pare nt 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 dibromobiphenyl s to rats, rabbits, and pigs suggests that the urine is an important route of excretion for polar metabo lites (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 locat ed 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 ar e also called biologically based tissue dosimetry models. PBPK models are increasingly used in risk a ssessments, 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 speci es (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.PBBs and PBDEs 267

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PBPK/PD models refine our unde rstanding of complex quantitativ e 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 ar e 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 subpopul ations within a species. The biological basis of PBPK models results in more meaningful extrapolati ons than those generated with the more conventional use of uncertainty factors.

The PBPK model for a chemical substance is deve loped 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 deve loped for a number of toxicologically important chemical substances, bot h volatile and nonvolatile (Krishnan and Andersen 1994; Leung 1993). PBPK models for a particular subs tance require estimates of the chemical substancespecific physicochemical parameters, and species-specif ic 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. Computer s 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 redu ces the magnitude of cumulative uncertainty. The adequacy of the model is, therefore, of great impor tance, and model validation is essential to the use of

PBPK models in risk assessment.

PBPK models improve the pharmacoki netic extrapolations used in ri sk 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 tissu e dose of chemicals in humans who are exposed to environmental levels (for ex ample, 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

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### Figure 5-5. Conceptual Representati on of a Physiologically Based Pharma cokinetic (PBPK) Model for a Hy pothetical Chemical Substance

Source: adapted from Krishnan et al. 1994

Note: This is a conceptual representation of a physiologi cally 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 skinPBBs and PBDEs 269

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Polybrominated Biphenyls. A PBPK model that incorporates tissue volume, affinity for PBBs, and rate of perfusion was developed to describe the distri bution and body burden of the major component of FireMaster PBB mixtures, 2,2',4,4',5,5'-hexabromo biphenyl (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 co rresponding 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 w ill be reduced proportionally. Congeners that cannot be readily metabolized (as is the case for BB 153) or ex creted 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 a nd depletion from all tissues of those congeners that can be cleared. In rats orally dosed daily with 14 C-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, an d clearance and rate constant parameters, human intake of 9.8 g of the congener from milk consum ption over a 230-day period would result in peak concentrations of 720 and 2.1 ppm in fat and blood, r espectively. 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 hum an 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 th e blood stream from the lungs, skin, or gastrointestinal tract is not known and little inform ation is available on how PBBs are distributed in thePBBs and PBDEs 270

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body. The available data indicate that the absorptio n 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 environmentall y contaminated blood, the distribution of PBBs among plasma, erythrocytes, mononucleocytes, and po lymorphonucleocytes 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 co nsumption 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 propertie s with other structurally similar polyhalogenated aromatic compounds, particular ly 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 particul ar, the oxygen bridge of the ether linkage in the diphenyl oxide molecule increases the distance betwee n 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 rotatio n 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 fo r 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 implicati ons not only for dioxin-type toxicities, which arePBBs and PBDEs 271

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mediated by the Ah (aryl hydrocarbon) receptor (AhR) pa thway, but also for non-dioxin-type effects. For example, Chen et al. (2001) found that the inductio n 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. Structur e-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 a nd 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/chemi cal properties that can influence the relative bioavailability, absorption, tissue accu mulation, 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 chlo rine also have implications for understanding the mechanism(s) of effects for nondioxin-like PBB c ongeners, 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 hydrocar bons are believed to share a common mechanism of action strongly related to similarities in their structur al 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-toxic ity relationships (Goldste in and Safe 1989; Safe 1990). Most of the studies providing this informa tion 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.PBBs and PBDEs 272

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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 extr ahepatic tissues of laboratory animals, mammalian cell cultures, and human organs and cell cultures (Gol dstein and Safe 1989). The structure-binding relationships for the coplanar 3,3',4,4',5-pent abromobiphenyl (BB 126), the monoortho substituted congener 2,3,3',4,4',5-hexabromobi phenyl (BB 156), and the diortho substituted analog 2,2’,5,5'-tetrabromobiphenyl (BB 52) were examined in rat liver cy tosol (Safe et al. 1985). At PBB concentrations 1,000-fold (10 μ M) 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 cy tosol (Millis et al. 1985). The results showed than the tetrabromobiphenyl was 10 times more effec tive than the hexabromobiphenyl in displacing radiolabeled 2,3,7,8-TCDD from the receptor. The ster eospecific nature of the bi nding (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 xe nobiotic metabolizing enzymes (cytochrome P-450- dependent monooxygenases) is we ll 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 CY P1A1 and CYP1A2 genes are induced by AhR agonists, such as 2,3,7,8-TCDD and MC, and the stru cture-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 exam ple, 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 patt ern of induction resembling that of MC (Robertson et al. 1982). Similar type experiments have show n that monoortho-bromo-substituted analogs of the coplanar PBBs, such as 2,3',4,4'-tetrabromobipheny l (BB 66), 2,3',4,4',5-pentabromobiphenyl (BB 118), and 2,3',4,4',5,5'-hexabromobiphenyl (BB 167), exhi bit 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 thirdPBBs and PBDEs 273

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group of PBB congeners, the diortho-bromo analogs of th e coplanar PBBs, resemble PB in their mode of induction (PB induces the CYP2B1 and CYP2B2 gen es). 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 so me dibromobiphenyls revealed that 4,4'-dibromobiphenyl (BB 15) resembled PB in its mode of in duction (Robertson et al. 198 2), whereas 2,2'-dibromobiphenyl (BB 4) had no significant effect on hepa tic 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 P BBs 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 e fficacious inducers of hepatic phase II metabolizing enzymes such as glutathione transferases, UDP glucur onyl 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 tran sferases while concomitantly depressing seleniumdependent glutathione peri oxidase activity, an important antioxidant enzyme in the liver (Schramm et al. 1985).

Many studies that examined structure-induction rela tionships 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 expe riments in rats revealed that of a series of PBB congeners, only MC-type inducers significantly decre ased 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 spleni c 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'-tetrabrom obiphenyl (weak PB-type) were compared in rats (Robertson et al. 1983a). Only BB 77 caused signifi cant reductions in growth rate and thymus size and marked depletion of lymphocytes from the thymic co rtex. Results from studies with FireMaster BP-6 revealed that the pattern of toxic responses and the ma gnitude of the responses attributed to this mixture are consistent with it being composed of both MC-t ype and PB-type congeners; the most toxic responses being attributed to the MC-type congeners (Akoso et al. 1982a, 1 982b; Dannan et al. 1982a, 1982b; Ecobichon et al. 1979; Parkinson et al. 1983; Render et al. 1982). These results suggest a correlationPBBs and PBDEs 274

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between immunological and hepatic effects and the ab ility to induce AHH activity and that the most toxic congeners are those that resemble the structural conf iguration of 2,3,7,8-TCDD. This relationship further supports the idea of a common receptor-mediated mechanism of action. Other PBB congeners ( ortho substituted) 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 lesi ons) was assessed in rats that were partially hepatectomized, initiated with diethylnitrosamine an d 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'-hexabromobiphe nyl [BB 153]) congeners s howed 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), whic h also indicates that other congeners are very effective as promoters, or possibly that the combin ation 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 cong eners may exert their effects indirectly by causing hepatotoxic (cytotoxic effects and necrosis), wh ereas the PB-type congeners may act as mitogens (stimulate cellular growth and division). In a sk in 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 interc ommunication assay (Kang et al. 1996) . Gap junctional intercellular intercommunication in normal human breast epithelia l 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 interce llular communication is generally rega rded 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 bindingPBBs and PBDEs 275

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affinity for inducers of benzo[a]pyrene hydroxylase such as 3-MC, ß-naphthofl avone, 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 ar omatic 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 redu ced sensitivity to more potent inducers such as 2,3,7,8-TCDD. Studies with PBBs have shown that tr eatment of C57BL/6J and DBA/2J with FireMaster BP-6 resulted in the induction of hepatic microsom al 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 mo re potent MC-type inducer than the BP-mixture, induced benzo[a]pyrene hydroxylase in both st rains of mice but did not induce aminopyrine N-demethylase in either strain of mice. Also, afte r 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 r esponses observed. It must be mentioned, however, that differences in the response between tissues, strain s, or species, do not exclusively indicate differences in receptor affinities, but most likely reflect the f act 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 fo llowed by an activation or transcription step and subsequent accumulation of occupied nuclear recept or 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 expressi on. Newly synthesized enzymes and macromolecules resulting from the pleiotropic response to the ligand-r eceptor complex are believed to be responsible for many of the effects caused by PBBs and other haloge nated 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 cy cle regulation (e.g., apoptosis, proliferation, and differentiation), morphogenesis, oxidative stress re sponse, and other processes, and is ultimately expressed as a diverse spectrum of well characterized toxic responses.

No studies were located regarding the mechanism of e ndocrine effects (thyroid toxicity, estrogenicity) of

PBBs.PBBs and PBDEs 276

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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 respon ses are very weak or negligible, particularly for

commercial PBDE mixtures and envi ronmentally relevant congeners.

For example, Meerts et al. (1998) indirectly exam ined the AhR-mediated (dioxin-like) properties of 17 PBDE congeners in a recombinant H4II rat hepato ma 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 (B DE 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 ra t 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 ma gnitude 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, includi ng 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. Over all, the tested PBDE congeners were at least 200,000 times less potent than TCDD for inducing AhR-medi ated 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 pe ntaBDE 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 bi nding affinity was only 2% that of 2,3,7,8-TCDD. No binding activity could be determined for the decaB DE 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 commer cial mixtures was subsequently more completely characterized, by determining whether they act as Ah R agonists or antagonists at sequential stages of the AhR signal transduction pathway leading to CYP1A1 in rat hepatocytes (Chen and Bunce 2001).PBBs and PBDEs 277

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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 or ders of magnitude greater than TCDD. These congeners showed additive behavior towards DRE binding with TC DD (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’-hex aBDE (BDE 153), and 2,2’,3,4,4’,5’,6-heptaBDE (BDE 183) were very weak activators of DRE bindi ng, and other congeners and the three commercial BDE mixtures were inactive. In particular, the e nvironmentally 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 si gnal transduction), and the commercial pentaBDE mixture had negligible EROD induction activity. The PBDE c ongeners that bound most strongly to the AhR were also the strongest inducers of CYP1A1 mRNA and CY P1A1 protein, indicating that the induction of CYP1A1 was AhR-mediated. Considering all end point s 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 th e 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 (Darne rud 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 ER OD 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 hepato cytes 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'-tri BDE (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, theirPBBs and PBDEs 278

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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 affini ty (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-t oxicity 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 followi ng 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 ge neral, this potency order paralleled their potencies as inducers of hepatic microsomal AHH and EROD. Wo rth 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 toxi city 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 thyroi d 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 (T 4 ) 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 T 4 , 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 T 4 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 T 4 synthesis arePBBs and PBDEs 279

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unlikely. PBDEs are hepatic microsomal enzyme in ducers, but there is little evidence that increased enzyme activity leads to greater clearance of th yroid 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 T 4 (i.e., the conjugation of T 4 with glucuronic acid). An indication that increased UDPGT activity may not be the

main mechanism for the reduced T 4 levels is provided by Hallgren et al. (2001), who found that exposure

to ≥ 18 mg/kg/day pentaBDE for 14 days caused serum T 4 reductions in both mice and rats with no effect on UDPGT activity in the mice, and increased UDPGT in th e rats only at higher dose levels. In contrast,

the decreases in serum T 4 correlated with the induction of microsomal phase I enzymes (EROD and MROD). As discussed below, increased microsomal enzy me activity could also increase the formation of hydroxylated PBDE metabolites that can bind to T 4 plasma transport proteins. This would serve to increase the number of occupied sites on T 4 -binding proteins and subsequently result in decreased serum

levels of T 4 ; however, this mechanism is not fully elucidated.

The possible interaction of PBDEs with T 4 binding to human transthyretin (TTR) was investigated in an

in vitro competitive binding assay (Meerts et al. 1998, 200 0). Testing of 17 congeners, ranging from dito heptaBDEs, showed that none of the parent compounds competed with T 4 for binding to human TTR. Incubation of the congeners with rat liver mi crosomes induced by PB (CYP2B enriched), β -napthoflavone (CYPIA enriched), or clofibrate (CYP4A3 enriched) indicated that metabolism is necessary to compete with T 4 -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 T 4 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 T 4 -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), althoug h 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-T 2 ), 3,3',5-triiodothyronine (T 3 ), and 3,3',5,5'-tetraiodothyronine (T 4 ), were also tested in the T 4 -TTR competition binding assay. The relative potencies showed that the T 4 -like (2,6-dibromo-4-[2,4,6-tribromophenoxy]phenol)

and T 3 -like (2-bromo-4-[2,4,6-tribromophenoxy]phenol) h ydroxylated PBDEs were 1.42- and 1.22-fold more potent, respectively, than T 4 , and the percentage competition at 500 nM exceeded that of the natural ligand; the T 2 -like hydroxylated PBDE (4-[2,4,6-tribromophenoxy]phenol) showed low affinity for TTRPBBs and PBDEs 280

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(0.41-fold less potent than T 4 ). Because the PBDEs were able to compete with T 4 -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 h ydroxylation 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 octa BDE products. For example, of the in vitro -generated metabolites displacing T 4 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) show ed 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 st ructural 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 structur al relationships with thyroid hormones. It is relevant to note that high concentrations of h ydroxy 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 T 4 and T 3 . None of the hydroxylated derivatives effectively competed with the thyroid hormones for binding to either recepto r (affinities were 4–>1,000 times less than for T 4 and

T 3 ). Because the tested congeners were the most likel y to have affinity for the thyroid hormone receptor, it was speculated that other hydroxylated PBDE cong eners will have even lo wer potential for receptor binding. DecaBDE (not hydroxylated) had no effect on thyroid hormone receptor-mediated transcriptional activation by T 3 in HeLaTRDR4-luc human cells; no ot her congeners were tested in this assay (Sakai et al. 2003).PBBs and PBDEs 281

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The extent that PBDEs affect circulating levels of T 4 or T 3 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 be ing, 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 T 4 transport to the brain in both rats and man (Blay et al. 1993; Sinjari et al. 1998). TTR does not transport

T 4 from the bloodstream to the brain, but rather is the main T 4 binding protein in cerebrospinal fluid (CSF) in both rats and humans. In the rat, T 4 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 T 4 to sites in rat choroid plexus following in vivo (but not in vitro ) exposure (Sinjari et al. 1998). Choroid plexus homogena te from rats that were orally treated with 6 or 18 mg/kg/day for 14 days showed T 4 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 T 4 , normal free T 4 , and has apparent good health (Palha et al. 1997, 2000). Interference with th e blood-choroid-plexus-CSF-TTR-mediated route of T 4 to the brain caused by the absence of TTR did not prod uce measurable features of hypothyroidism (Palha et al. 2000).

Neurobehavioral developmental altera tions have been induced in mice th at 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 cha nges in hormone levels and other thyroid end points in animal models. Some studies sugg est that the effects might be related to alterations in cholinergic functions. For example, neonatal e xposure 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 expo sure to BDE 99 (single 8 mg/kg oral dose at age 5 months) also affected behavior in mice; the ch ange was not seen in mice only exposed to BDE 99 asPBBs and PBDEs 282

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adults or mice only exposed to nicotine as neonat es (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 mi xture DE-71 or tetra congener BDE 47 stimulated arachidonic acid release in rat cerebellar gr anule neurons; this effect was not seen with the commercial octaBDE product DE-79 (Kodavanti 2003; K odavanti 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 A 2 . In vitro exposure to penta mixture DE-71and tetra congener BDE 47 also caused translocation of protein kinase C, as indi cated by increased phorbol ester binding; octaBDE mixture DE-79 did not induce this effect (Kodava nti 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 (Kodavan ti 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 mo re toxic than octa and deca congeners in inducing

cell death and free radical formation in cerebella r granule cells (Reistad et al. 2002). The in vitro uptake of the neurotransmitter dopamine into rat brain synapt ic 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 3 H-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 sever al 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 hydrox ylated PBDEs, tribromophenoxy)phenol, 2-bromo-PBBs and PBDEs 283

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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 T 2 (3,5-diiodothyronine), T 3 (3,3’,5-triiodothyronine), and T 4 (3,3’,5,5’-tetraiodothyronine), respectively. Eleven of 17 PBDE congeners showed estrogenic activity (dose-dependent luciferase inducti on) 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 EC 50 values that were 250,000–390,000 times less potent than 17 β -estradiol (E 2 ). In contrast, the T 3 - and T 2 -like hydroxylated PBDEs showed estrogenic potencies exceeding that of E 2 (no estrogenic activity was induced by the T 4 -like hydroxylated PBDE). Antiestrogeni c potencies were determined in the ERCALUX assay by treating T47D.Luc cells with th e PBDEs and hydroxylated PBDEs in the presence of

E 2 . 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 E 2 -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-tribromo phenoxy)phenol, were also tested for estrogenicity

in ER α -specific and ER β -specific human embryonic kidney cell lines (293-ER α -Luc and ER β s-Luc cells, respectively). The hydroxylated PBDE was potent in the ER α -specific cells (maximum luciferase

induction similar to E 2 ) and also showed activity in the ER β -specific cells (maximum 50% induction compared to E 2 ), whereas the ER α - and ER β -specific cell lines were less responsive to 2,4,6-triBDE (BDE 30) (34.2 and 7.8% induction compared to E 2 ) and 2,2',4,4',6-pentaBDE ( ≈ 20 and <2% relative induction). These results indicate that pure and hydr oxylated congeners of PBDEs can be agonists of

both ER α and ER β 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 hydrox yl 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 mi xtures 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 refl ect multiple exposure pathways and congener-specific elimination and thus, in general, represent stead y state body burdens that do not match the congener profiles in the original exposure sources. For exam ple, profiles of PBB and PBDE congeners in human milk do not resemble the pattern of any of the comme rcial 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 pre dominate 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 inters pecies 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 e ffects of PBDEs on circulating levels of thyroid hormones. This difference is thought to derive fro m the rat thyroid having a smaller store of iodinated thyroglobulin that is more easily depleted when the av ailability of iodide is limited, and from a more rapid

clearance of T 4 from the rat circulation; the latter resulting from rats not having a high affinity binding protein for T 4 in serum analogous to thyroid-binding globulin (TBG) in humans (Capen 1997). If the production of T 4 and T 3 is impaired sufficiently to deplete the thyroid of stored iodinated thyroglobulin, the thyroid cannot produce or secrete amounts of T 4 and T 3 needed to support physiological demands, circulating levels of T 4 (free T 4 ) and T 3 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-T 4 is bound to TGB, and TTR and albumin bind only 11 and 15%, respectively, of the total (Schussler 2000). In cont rast to most mammals, the rat utilizes TTR as the major

T 4 plasma binding protein; approximately 75% of T 4 in rat serum is bound to TTR and only 25% to albumin. Both circulating T 3 and T 4 are highly protein bound with only a small fraction of their total present as free hormone, and this high degree of protei n binding serves to maintain equilibrium between the extracellular and intracellular pools of these horm ones (O’Connor et al. 1999). The increased

metabolic clearance of serum T 4 is thought to involve the induction of hepatic microsomal enzymes. Although it is well documented that PBDEs are microsom al enzyme inducers, there is little evidence that induction of hepatic enzymes leads to greater clea rance of thyroid hormones, indicating that it isPBBs and PBDEs 285

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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 human s 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 poten tial 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 gesta tion prior to the onset of fetal hormone production.

5.6 TOXICITIES MEDIATED THROUGH THE NEUROENDOCRINE AXIS

Recently, attention has focused on th e 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 referre d to as endocrine disruptors. However, appropriate terminology to describe such effects remains contr oversial. The terminology endocrine disruptors, initially used by Colborn and Clement (1992), was al so used in 1996 when Congress mandated the EPA to develop a screening program fo r “...certain substances [which] may have an effect produced by a naturally occurring estrogen, or other such endocrine effect[s]...”. To meet th is mandate, EPA convened a panel called the Endocrine Disruptors Screening a nd Testing Advisory Committee (EDSTAC), and in 1998, the EDSTAC completed its deliberations and ma de recommendations to EPA concerning endocrine disruptors. In 1999, the National Academy of Scien ces released a report that referred to these same types of chemicals as hormonally active agents. The termi nology endocrine modulators has also been used to convey the fact that effects caused by such chemic als 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 wildlif e. However, others think that endocrine-active chemicals do not pose a significant health risk, particular ly in view of the fact that hormone mimics exist in the natural environment. Examples of natura l 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 estr ogen. Although the public health significance and descriptive terminology of substances capable of affe cting the endocrine system remains controversial, scientists agree that these chemicals may affect the s ynthesis, secretion, transport, binding, action, or elimination of natural hormones in the body respons ible 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 toPBBs and PBDEs 286

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be involved in inducing breast, testicular, and pros tate 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 e ffects on humans and wildlife (Colborn et al. 1993; Crisp et al. 1998; Daston et al. 1997; Safe and Zachar ewski 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 ma ny target tissues, including female and male reproductive systems, such as mammary gland, uterus, va gina, ovary, testes, epididymis, and prostate. In addition, there is evidence that some of these envi ronmentally-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 estr ogenic 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 T 4 and T 3 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 FireM aster 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 a nd 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 T 4 , and increased thyroid antimicrosomal antibody titers (Bahn et al. 1980).

Serum levels of the adrenal hormones corticoste rone B, dehydroepiandr osterone, 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 corticoste rone levels and adrenal weight di d not change in rats exposed to

≤ 6 mg/kg/day of an unspecified PBB mixture for a shor ter duration of 20 days (Castracane et al. 1982).PBBs and PBDEs 287

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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; La mbrecht et al. 1978). A corresponding decrease in serum levels of progesterone suggests that the repr oductive effects in the monkeys are related to PBBinduced endocrine imbalance. Implantation was comple tely 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 F 1 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 feed ing (Blanck et al. 2000b; Thomas et al. 2001). Determinants of PBB serum decay were investigat ed in women who had a mean initial PBB level of 20.9 ppb, a mean time between the first and last meas urement of 4.2 years, and at least two measurements taken when they were not pregnant (Blanck et al. 200 0b). The median PBB half-life was estimated to be 13.5 years. Subject-specific decay rates were regress ed 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 resu lts of this study is provided in Section 5.8.1. Thomas et al. (2001) found no relationship between se rum 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 leve l 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 va riable 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 significan tly associated with serum PBB levels. Additional information on the design and results of this study is provided in Section 5.2.2.5.PBBs and PBDEs 288

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The issue of breast cancer has received attention foll owing reports of high levels of organochlorine compounds in breast cancer patients. A relationshi p 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 ca ses, 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 envir onmental 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 thyroi d 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 PB DE congeners showed estrogenic activity (dosedependent luciferase induction) in the ER-CALUX as say with T47D.Luc cells, although the most potent PBDE congeners (2,2',4,4',6-pentaBDE > 2,4,4',6-tet raBDE [BDE 75] > 2,2',4,6'-tetraBDE [BDE 51] > 2,4,6-triBDE [BDE 30] > 2,3',4,4',6-pentaBDE [BDE 119]) had EC 50 values that were 250,000– 390,000 times less potent than that of the natural ligand, 17 β -estradiol (E 2 ). In contrast, two of the hydroxylated PBDEs, 4-(2,4,6-tribromophenoxy)phe nol, 2-bromo-4-(2,4,6-tribromophenoxy)phenol (which have bromine substitution pattern s similar to the thyroid hormones T 2 [3,5-diiodothyronine] and

T 3 [3,3’,5-triiodothyronine], respectively) had estrogenic potencies exceeding that of E 2 . 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 ER α -specific and

ER β -specific human embryonic kidney cell lines (293-ER α -Luc and ER β s-Luc cells, respectively). The hydroxylated PBDE was potent in the ER α -specific cells (maximum luciferase induction similar to E 2 ) and also showed activity in the ER β -specific cells (maximum 50% induction compared to E 2 ), whereas

the ER α - and ER β -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 ER α and ER β receptors and that metabolism to hydroxylated PBDEs may increase estr ogenic potency. However, no estrogenic activity was demonstrated by the predominant congeners in bi ological samples, 2,2’,4,4’-tetraBDE (BDE 47) andPBBs and PBDEs 289

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2,2’,4,4’,5-pentaBDE (BDE 99). The hydroxy-PBDE derivatives that showed minimal activity were

para -substituted with a hydrox yl 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 estrogeni city 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 T 4 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 T 4 for binding to human TTR. Incubation of the congeners with rat liv er 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 iden tified) that were able to displace T 4 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-T 2 ), 3,3',5-triiodothyronine (T 3 ), and 3,3',5,5'-tetraiodothyronine (T 4 ) showed that the T 4 -like (2,6-dibromo-4-[2,4,6-tribromophenoxy]phenol) and T 3 -like (2-bromo-4-[2,4,6-tribromophenoxy]phenol) hydroxylated PBDEs were 1.42- and 1.22-fold more potent, respectively, than T 4 ; the T 2 -like hydroxylated PBDE (4-[2,4,6-tribromophenoxy]phenol) showed low affinity fo r TTR (0.41-fold less potent than T 4 ). Studies with hydroxylated derivatives of PBBs similarly showed that the c ongener patterns most closely resembling the diiodophenolic ring of thyroxine had the highest TTR bi nding activity (Chauhan et al. 2000; McKinney et al. 1987; Rickenbacher et al. 1986).

Because PBDEs were able to compete with T 4 -TTR binding only after me tabolic conversion, and considering that thyroid hormones are also hydroxy-ha logenated 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 T 4 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) show ed that BDE 47 is well absorbed (~95%), poorly metabolized (only 3% of the amount excreted in f eces was in the form of metabolites), and occurs inPBBs and PBDEs 290

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blood and other tissues mainly as the parent compound. Additionally, the three pure hydroxylated PBDEs synthesized by Meerts et al. (1998, 2000) for their st ructural 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 theoreti cally show the highest structural similarity to T 4 and T 3 . 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 T 4 and T 3 ). Because the tested congeners were the most likely to have affinity for the thyroi d 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 T 4 levels and follicular cell hyperplasia have been consistently observed in rats and mice orally exposed to PBDEs. Acut e-duration studies showed decreases in serum T 4 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 studi es include thyroid hyperplasia in rats exposed to

≥ 8 mg/kg/day octaBDE for 30 days (Norris et al. 1973, 1975a, 1975b) and reduced serum T 4 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 T 4 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 cau sed 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 (T 4 ) levels with no accompanying changes in serum TSH (Darnerud and Sinjari 1996; FowlesPBBs and PBDEs 291

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et al. 1994; Hallgren and Darnerud 1998;WIL Research Laboratories 1984; Zhou et al. 2001, 2002), and decaBDE appears to be much less pot ent than the lower brominated mixtures. Considering these data, the

structural resemblance of some PBDEs to T 4 , and information from binding studies of individual congeners as summarized above, it is hypothesi zed 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 T 4 or T 3 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 ab sence of thyroid binding globulin (TBG) in rats (Capen 1997). Transthyretin (TTR) is the major thyroid horm one binding protein in rats, whereas TBG is the main binding protein in man and most other mamma ls. However, although TTR is a minor thyroid hormone binding protein in humans, it is the principal protein involved in T 4 transport to the brain in both rats and man (Blay et al. 1993; Sinjari et al. 1998). TTR does not transport T 4 from the bloodstream to the brain, but rather is the main T 4 binding protein in cerebrospinal fluid (CSF) in rats and humans. In the

rat, T 4 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 T 4 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 bi ological 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 mate rnal exposure during gestation and lactation.

Relevant animal and in vitro models are also discussed.

Children are not small adults. They differ from a dults in their exposures and may differ in their susceptibility to hazardous chemicals. Children’s uni que physiology and behavior can influence the extent of their exposure. Exposures of children ar e discussed in Section 8.6 Exposures of Children.

Children sometimes differ from adults in their susceptib ility 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 th e relationship may change with developmental agePBBs and PBDEs 292

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(Guzelian et al. 1992; NRC 1993). Vulnerability ofte n depends on developmental stage. There are critical periods of structural and functional devel opment during both prenatal and postnatal life and a particular structure or function will be most sensitiv e to disruption during its critical period(s). Damage may not be evident until a later stage of developmen t. There are often differences in pharmacokinetics

and metabolism between child ren and adults. For example, absorption may be different in neonates because of the immaturity of their g astrointestinal tract and their larg er skin surface area in proportion to body weight (Morselli et al. 1980; NRC 1993); the gastroin testinal 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 ex tracellular water and their brains and livers are proportionately larger (Altman and Dittmer 1974; Fomon 1966; Fom on 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 ma y be higher or lower than those of adults, and sometimes unique enzymes may exist at particular de velopmental stages (Komori et al. 1990; Leeder and Kearns 1997; NRC 1993; Vieira et al. 1996). Whethe r differences in xenobiotic metabolism make the child more or less susceptible also depends on whethe r the relevant enzymes are involved in activation of the parent compound to its toxic form or in detoxifica tion. There may also be differences in excretion, particularly in newborns who all ha ve a low glomerular filtration rate and have not developed efficient tubular secretion and resorption capacities (Altman a nd Dittmer 1974; NRC 1993; West et al. 1948). Children and adults may differ in their capacity to repa ir damage from chemical insults. Children also have a longer remaining lifetime in which to express da mage 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 di fference might be somewhat counterbalanced by their alveoli being less developed, which results in a di sproportionately smaller surface area for alveolar absorption (NRC 1993).

Children are exposed to PBBs and PBDEs in the sam e manner as the general population, primarily via consumption of contaminated foods. Exposure also ma y 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.PBBs and PBDEs 293

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Placental transfer, although it may be limited in absolu te amounts, is a concern because of possible effects of PBBs and PBDEs on sensitive immature tissues, or gans, and systems, with potentially serious longlasting consequences. Transfer of PBBs and PBDEs vi a 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 child ren 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 prof iles 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 childre n or nutritional factors that may influence the absorption of PBBs.

No biomarkers of exposure or effects for PBBs or PB DEs 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 P BBs or PBDEs with other chemicals in children or adults. No information was located regarding pedi atric-specific methods for reducing peak absorption following exposure to PBBs or PBDEs, reducing bo dy 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 ra ther 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 Spei zer 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.PBBs and PBDEs 294

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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 c onsistent or marked abnormalities. No significant abnormalities were found by physical and neuropsychological examination of 33 of these exposedchildren 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 fre quent 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 Child ren'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 perfo rmances were within the normal range in all areas assessed (Schwartz and Rae 1983). Due mainly to the sm all data set and the inco nsistency of the results, the available data do not conclusively es tablish or eliminate the possibility that in utero and early infancy exposure to PBBs might adversely affect the development of human children.

Neurobehavioral alterations have b een observed in animals following gestational and lactational exposure to PBBs. Performance deficits in tests of operant beha vior were seen in 6-month-old offspring of rats that were exposed to ≥ 0.2 mg/kg/day of FireMaster BP-6 by ga vage from day 6 of gestation until weaning (Henck and Rech 1986), but not in 75-da y-old offspring of rats exposed to ≥ 0.5 mg/kg/day for 4 weeks prior to mating (Geller et al. 1985). Effects on acqui sition of forward locomotion, cliff avoidance, cage emergence, and open-field activity were f ound 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 dur ing gestation and through weaning showed altered negative geotaxis and avoida nce 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 19 77, 1979; Groce and Kimbrough 1984) and mice (Luster et al. 1980) after oral exposur e during gestation and in monkeys after exposure before conception and during pregnancy (Allen et al. 1979; Lambrecht et al. 1978). StructuralPBBs and PBDEs 295

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malformations in fetuses, including cleft palate, we re also observed in rats (Beaudoin 1977) and mice (Corbett et al. 1975) after exposure to these PBBs dur ing gestation. Increased incidences of fetuses with extra ribs were found in a study of rats orally ex posed to a commercial octabromobiphenyl mixture during gestation (Waritz et al. 1977), although decabrom obiphenyl 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 offspr ing of rats and mice after exposure during gestation and/or lactation (Corbett et al. 1975; Groce a nd Kimbrough 1984; McCormack et al. 1981, 1982c; Meserve et al. 1992). Liver effects, including in creased liver weight and hepatic cytochrome P-450 enzymic activity, hepatocyte enlargement, vacuolizati on, and/or other degenerati ve 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 T 4 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 injec tion of corticotropin-releasing factor or adrenal stimulation by an injection of adrenocorticot ropic hormone. Serum concentrations of T 3 and T 4 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 offs pring are consistent with evidence that the thyroid gland is an unequivocal target of PBBs in adult animals. A sp ectrum of thyroid effects, ranging from decreases in

serum T 4 and T 3 levels to histological and ultrastructural ch anges in the follicles, has been documented in adult rats orally exposed to PBBs (mainly FireM aster 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 in cluded increased serum FSH, low or borderline low

serum T 4 , and increased thyroid antimicrosomal antibody titers (Bahn et al. 1980).

No information was located on possible immunological e ffects 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 co ntamination episode have been reported in some studies (Bekesi et al. 1978; Roboz et al. 1985), but othe r investigations were not able to confirm these findings (Landrigan et al. 1979; Silva et al. 1979; Stross et al. 1981). No correlation can be establishedPBBs and PBDEs 296

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between altered immune parameters and serum PBB le vels 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 inte rmediate-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 measur ed 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). PB B levels in breast milk of five women from exposed farms ranged from 0.21–92.7 ppm (Cordle et al. 1978; Hu mphrey and Hayner 1976). On a lipid basis, the

ratio of PBBs in breast milk to maternal serum w as 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 investigat ed 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 amount s 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 predicto r 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 deliver y of 9.4 ppb, a mean duration of breast-feeding as the main source of nutrition for 2.6 months, and a me an total duration of breast-feeding of 4.1 months. Exposure was treated as a categorical variable by di viding the women into groups of low (reference)PBBs and PBDEs 297

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exposure ( ≤ 1 ppb), moderate exposure (>1– ≤ 7 ppb), and high exposure (>7 ppb). Three outcomes of interest were analyzed: (1) the decision to breast f eed (yes/no), (2) the duration (months) of breast-feeding as the main source of nutrition, and (3) the total durat ion (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. Thyroi d 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, migra tion, and differentiation du ring development, and maintenance of normal levels is essential to normal gr owth 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 th e 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 in fants (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 th yroid 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 al ready a widely accepted and legislatively mandated practice (LaFranchi 1999; Landenson et al. 2000); newbor ns 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).PBBs and PBDEs 298

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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 comme rcial mixtures and single congeners in adult and

neonatal rats and mice. The main effects in animals are reduced serum T 4 hormone levels and follicular cell hyperplasia, with no accompanying changes in serum T 3 or TSH levels. In adult animals, acuteduration oral exposure caused decreases in serum T 4 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 Darn erud 1998; Hallgren et al. 2001; Zhou et al. 2001).

Effects observed in intermediate-duration oral studi es included thyroid hyperplasia in rats exposed to

≥ 8 mg/kg/day octaBDE for 30 days (Norris et al. 1973, 1975a, 1975b) and reduced serum T 4 in rats exposed to ≥ 10 mg/kg/day pentaBDE for 90 days (WIL Research Laboratories 1984).

Other animal studies have shown redu ced serum levels of thyroid hormones in offspring of rats exposed to lower brominated PBDEs during gestation and lactati on, as well as in rats orally exposed as weanlings. Exposure to pentaBDE from GD 6 through the end of lactation caused serum T 4 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-ol d) 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 T 4 and T 3 levels at ≥ 10 and ≥ 60 mg/kg/day, respectively (Zhou et al. 2001). Similar exposure to pentaBDE caused serum T 4 and T 3 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 T 4 hormone levels and follicular cell hyperplasi a in adult rats and mice confirming that the thyroid is a sensitive target of octaBDE and pent aBDE 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 ho rmone homeostasis can cause neurodevelopmental effects, little specific information is currently available on the potential neur otoxic effects of PBDEs. Data are mainly limited

to the results of three behavioral tests in animal s 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 mi ce treated at either 3 or 10 days of age, but not at 19 days of age, suggesting that there is a criti cal 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,PBBs and PBDEs 299

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octaBDE, and pentaBDE products have been screen ed for neurotoxicity using comprehensive test batteries that typically include functional observationa l, 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 opportun ities 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 neurodevelo pmental 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 T 4 transport to the brain in both rats and man (Blay et al. 1993; Sinjari et al. 1998). TTR does not transport T 4 from the bloodstream to the brain,

but rather is the main T 4 binding protein in cerebrospinal fluid (CSF) in rats and humans. In the rat, T 4 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 co nsequent increased metabolic formation of hydroxymetabolites, but humans are not particularly sensitive to this effect. Additionally, the animal studies in

which neurobehavioral alterations we re 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 thyroi d and neurodevelopmental effects by decaBDE is low in comparison to commercial penta- and octaBDE mixtures due to to xicokinetic and toxicity differences. The lower brominated BDEs preferentially accumulate in the bod y due to their partitioning and retention in lipidrich tissues and lower rates of metabolism and elimin ation 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 pent aBDEs appear to be relatively well absorbed, whereas the fully brominated decaBDE, a large poorly so luble molecule, is very poorly absorbed ( ≈ 1% or less ofPBBs and PBDEs 300

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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 mi xtures. 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 pe ntaBDE mixtures caused decreases in serum T 4 and T 3 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 do se 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 immunos uppressive potential of PBDEs in young animals, and data on immune function in adult animals are mainly li mited to findings in acute-duration oral studies of relatively high doses of pentaBDE. The plaque-formi ng 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 Thuva nder 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 a nd 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 fema les) 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 pr ovided opportunities for the development of histopathology in immune system tissues and releva nt clinical signs, no comprehensive immunotoxicity evaluations of PBDEs have been performed.PBBs and PBDEs 301

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Information on the reproductive toxicity of PBDEs main ly consists of a one-generation study of decaBDE in rats that found no exposure-related functional e ffects (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 t oxicity 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 toxi city 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 ha ve 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 est rogenic 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 (Meer ts et al. 2001). Eleven of 17 PBDE congeners showed estrogenic activity (dose-dependent luci ferase induction) in the ER-CALUX assay with T47D.Luc cells, although the most potent PBDE conge ners (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 EC 50 values that were 250,000–390,000 times less potent than that of the natural ligand, 17 β -estradiol (E 2 ). In contrast, two of the hydroxylated PBDEs, 4-(2,4,6-tribromophenoxy)phenol, 2-bromo-4-(2,4,6-tribromophenoxy)phenol (which ha ve bromine substitution patterns similar to the thyroid hormones T 2 [3,5-diiodothyronine] and T 3 [3,3’,5-triiodothyronine], respectively) had estrogenic potencies exceeding that of E 2 . 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 ER α -specific and ER β -specific human embryonic kidney cell lines (293-ER α -Luc and ER β s-Luc cells, respectively). The hydroxylated PBDE was potent in the ER α

specific cells (maximum luciferase induction similar to E 2 ) and also showed activity in the ER β -specific cells (maximum 50% induction compared to E 2 ), whereas the ER α - and ER β -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 ER α and ER β receptors and that metabolism to hydroxylated PBDEs may increase estrogenic potency. However, no estrogenic activity w as 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-PBBs and PBDEs 302

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PBDE derivatives that showed minimal activity were para -substituted with a hydrox yl 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, particular ly commercial pentaBDE mixtures whose constituent tetra- to hexaBDE congeners are highly bioaccumulati ve, 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 av ailable for transfer to breast milk is low due to poor gastrointestinal absorption and rapid elimin ation in the feces (Hardy 2002b). The increasing temporal trend for PBDEs in human tissues is illustrate d 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 ch ildbirth. No PBPK mode ls have been developed for PBDEs that could be used to quantitatively predic t transfer of PBDEs via breast milk or across the placenta.

Assessments of potential health risks to children we re 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, 200 3b). Based on evaluations of toxicity and potential exposure data, the reports concluded that n one 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 re flecting peer review findings and EPA conclusions are not yet available (EPA 2004a, 2004b, 2004c).PBBs and PBDEs 303

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5.8 BIOMARKERS OF EXPOSURE AND EFFECT

Biomarkers are broadly defined as indicators signaling ev ents 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 interpre tation of biomarkers, impl ementation 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. Ho wever, several factors can confound the use and interpretation of biomarkers of e xposure. The substance being measur ed 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., dura tion and route of exposure), the s ubstance 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 bi ochemical, 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 definiti on 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 su ch as increased blood pressure or decreased lung capacity. Note that these markers are not often subs tance 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 inhe rent or acquired limitation of an organism's ability to respond to the challenge of exposure to a specific xenobi otic substance. It can be an intrinsic genetic or other characteristic or a preexisting disease that result s in an increase in absorb ed dose, a decrease in the biologically effective dose, or a target tissue res ponse. If biomarkers of susceptibility exist, they are discussed in Section 5.10 “Populations That Are Unusually Susceptible”.PBBs and PBDEs 304

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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 e xposure to PBBs. Because they are lipophilic and have long half-lives, certain PBB congeners preferen tially accumulate in lipid-rich tissues, especially adipose, and are present in serum and human milk. Bo th serum and adipose PBB levels are indicators of exposure, but monitoring PBBs simultaneously in sam ples 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 fo llowing 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 r esults were included reflects the higher fat content of adipose compared to serum and the lipophilici ty 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 i ndicate 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 consta nt parameters, it was predicted th at 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 excre tion rate in a lean individual was estimated at

10.2 μ g/day; overweight individuals ha d an excretion rate of 4.1 μ g/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 t 1/2 increases with increasing fat content. These predictions point to th e 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 thePBBs and PBDEs 305

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same exposure (Tuey and Matthews 1980). The assump tions on which the predictions are based do not reflect possible differences in fat and lean subject s 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-lif e of 6.5 years predicted by Tuey and Matthews (1980) is shorter than half-life values determined using ser a 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 leve ls to fall below a detection limit of 1 ppb.

Determinants of PBB serum decay were inves tigated in a group of 380 Michigan women ( ≥ 16 years of age) who had an initial PBB level of at least 2 ppb an d at least two measurements taken when they were not pregnant (Blanck et al. 2000b). The mean initia l PBB level was 20.9 ppb (standard deviation 78.7),

and the mean time between the first and last measureme nt 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 amount s 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 we re 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, w ith higher initial PBB levels; women with initial PBB levels of <10 and >10 p pb 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 meas urement was also associated with slower decay, but the effect was of borderline statistical significan ce (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.PBBs and PBDEs 306

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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 So vocool 1982). Chemical workers involved in the PBB manufacturing process had a median adipose P BB 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 elim ination compared to decaDBE (i.e., related to each molecule’s specific structure and not solely due to lip ophilicity). Serum, adipo se, 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 widespr ead production and use of PBDEs, particularly commercial pentaBDE mixtures whose tetra- to he xaBDE congeners are highly bioaccumulative, as well as the time lag from current-year usage to exposure vi a the food chain, it is highly likely that tissue concentrations among the general popul ation 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. 19 99). 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 dura tion 28 days, range 21–35 days) indicate that the higher brominated congeners have shorter half-lives th an 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),PBBs and PBDEs 307

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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 informatio n 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 T 4 , and increased thyroid antimicrosomal antibody titers (Bahn et al. 1980). A spectrum of th yroid effects has been observed in exposed rats, ranging from decreases in serum levels of serum T 4 and T 3 to histological and ultrastructural changes (Akoso et al. 1982b; Allen-Rowlands et al. 1981; Byrn e 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 T 4 and/or other thyroid hormones are potential biomarkers of effects for PBBs. Additiona l studies could better characterize thyroid effects of PBBs in humans and develop specific correlations between levels and du ration of exposure and

alterations in serum T 4 and T 3 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 to ol 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 thePBBs and PBDEs 308

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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 mo nitored by the CBT, and 7-N-demethylase activity was monitored by the caffeine urinary metabolite ra tio (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 (r 2 =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. Ge nerally, smokers have higher CBT values than nonsmokers due to the presence of polycyclic aroma tic hydrocarbons (PAH) in tobacco smoke, which induce CYP1A (Kotake et al. 1982).

Many reports have been published regarding possi ble associations between PBB exposure and adverse health effects in populations from the state of Michig an. An early study compared the health status of people on quarantined farms with people in nonqua rantined farms in the same area (Humphrey and Hayner 1975). The results show ed 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 (Landr igan 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-select ed 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 Mi chigan contamination episode. Furthermore, the prevalence of the reported symptoms had no consistent re lationship 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 biom arkers 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.PBBs and PBDEs 309

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Polybrominated Diphenyl Ethers. Biomarkers of effects for PBDEs are likely to be common to the general class of halogenated ar omatic hydrocarbons, rather than specific for PBDEs, because PBBs, PCBs, and other structurally similar chemicals cause ge nerally 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 produc ts caused thyroid changes in rats and mice, particularly reduced serum T 4 levels (Darnerud and Sinjari 199 6; 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 differenc es. In particular, lower bromin ated BDE congeners preferentially accumulate in the body compared to decaBDE, apparen tly 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 an imals following lifetime exposure to dose levels several orders of magnitude higher than effect leve ls of penta- and octaBDE in acute-duration studies (NTP 1986; Zhou et al. 2001).

Serum T 4 may not be a reliable biomarker in humans because the extent that PBDEs affect circulating

levels of T 4 (or T 3 ) possibly varies with species. As discussed in Section 5.5.3, Animal-to-Human Extrapolations, humans are generally regarded as less sen sitive than rats to thyroid effects of PBDEs due to differences in the binding of T 4 to serum transport proteins. In particular, in humans, the binding of

T 4 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 T 4 due to increased availability of the hormone for metabolism and

elimination. The clearance of T 4 in rats is enhanced by the indu ction hepatic microsomal enzymes, particularly UDP-glucuronyl transferase, which catalyzes the conjugation of free T 4 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. Howeve r, because PBDE-induced decreases in T 4 appear to involve multiple mechanisms in addition to increased thyr oid hormone metabolism (Zhou et al. 2001), serum T 4 level is a potential biomarker of effects for lowe r 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 T 4 . 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 legisla tively mandated practice (LaFranchi 1999; Landenson et al. 2000).PBBs and PBDEs 310

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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 H ook 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 weight s 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 weight s and by levels of SGPT and/or SGOT. In most cases, exposure to PBBs alone did not affect the para meters of nephrotoxicity in animals. However, exposure to PBBs alone usually caused increased relati ve liver weights and elevated levels of SGOT and SGPT. Pre-exposure to PBBs increased the hepa totoxicity and nephrotoxicity of chloroform (Ahmadizadeh et al. 1984; Kluwe and Hook 1978; Kluw e et al. 1978) and carbon tetrachloride (Kluwe et al. 1979, 1982) and the nephrotoxicity of trichloroeth ene 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 poten tiated the lethality of chloroform, carbon tetrachloride, and 1,1,2-trichloroethane by decreasing the LD 50 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 w ith PBBs in dietary studies increased the LD 50 value of 1,2-dibromo-3-chloropropane (DBCP) but had no effect on the LD 50 value of 1,2-dibromoethane (EDB) (Kluwe et al. 1981). Also, PBBs were found to reverse the de pletion 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 we re co-exposed to diets containing PBB and mirex, photomirex, or kepone, compared with toxicity elic ited by each of these substances alone (Chu et al. 1980).PBBs and PBDEs 311

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Polybrominated Diphenyl Ethers. No specific information was locat ed regarding interactions between PBDEs and other chemicals. PBDEs are a group of 209 congeners that display biological activity involving different potential mechanisms, and share so me toxicological properties with structurally similar polyhalogenated aromatic compounds, particularly PBBs, PCBs, polychlorinated- and polybrominated dibenzo- p -dioxins (PCDDs and PCBDs), and chlo rinated and brominated dibenzofurans (PCDFs and PBDFs). However, although these chemical s 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 (m ediated 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 effect s 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 ox ygenases, induction of changes in brain dopamine levels, and disruption of calcium homeostasis), or possibly both Ah-receptor-dep endent 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 toxici ty of PBDEs. For example, lower brominated PBDE mixtures and congeners are inducers of hepatic micr osomal 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 AR E UNUSUALLY SUSCEPTIBLE

A susceptible population will exhibit a different or e nhanced 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

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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 a nd 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, th e 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 huma ns 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 μ g/day; overweight individuals ha d an excretion rate of 4.1 μ g/day. The

cumulative excretion was 51% of the dose in lean s ubjects 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 con centrations 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 fatte r individuals. Because of this phenomenon, a sudden reduction in body fat, such as that which could occu r during dieting, may cause a redistribution of PBBs to potential target organs, which would also increas e the potential for adverse health effects to such

individuals.

Pregnant women and developing in fants 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 fe tus is dependent on maternal thyroid hormones at least until the fetal thyroid begins to produce T 4 and T 3 , which occurs in humans at approximately 16– 20 weeks of gestation (Zoeller and Crofton 2000). Th e potential of lower brominated PBDEs to disrupt maternal, fetal, and newborn thyr oid hormone levels is demonstrat ed by the Zhou et al. (2001, 2002)PBBs and PBDEs 313

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studies of pentaBDE (technical mixture) in rats, as discussed in Section 5.2.2.2 (Endocrine Effects subsection). The potential for induction of thyroid an d 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., lith ium), 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. Similarl y, people with compromised function of other organs, such as those with liver or kidney diseases (e.g., liver cirrhosis or hepatitis B), c ould 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 me chanistic data as discussed Section 5.5.3 (Animal-toHuman Extrapolations).

5.11 METHODS FOR REDUCING TOXIC EFFECTS

This section will describe clinical practice and resear ch 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 contro l 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 met hods that would apply to any persistent, lipophilic chemical, and have not been tested for efficacy, indica ting that they might not be effective in reducing the

toxic effects of PBBs and PBDEs. There is no indica tion 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-threateni ng acute exposure would occur are unlikely, although accidental or intentional inges tion of the commercial products is a conceivable

concern. The relevance of the methods to co mmon 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

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5.11.1 Reducing Peak Absorption Following Exposure

Ingested PBBs and PBDEs are absorbed by the g astrointestinal tract of humans and animals (see Section 5.4). Although there are no specific recommenda tions 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 relate d 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, e ither aqueous or mixed with a saline cathartic or sorbitol, is frequently recommended to decrease the gast rointestinal 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 gastroint estinal 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 ab sorbed by the lungs,skin, and/or orally by handto-mouth contact. Although no specific methods to re duce 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 th e dermal absorption of PCBs (HSDB 1992). Studies with monkeys showed that soap and water was as ef fective 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). Se veral methods to enhance the elimination of PBBs from the body have been examined in animals and are a pplicable to PBDEs, although the relevance of the methods is questionable because it is unclear whether current tissue levels are a health concern for thePBBs and PBDEs 315

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general population. Methods for increasing the elimin ation of these chemicals include the restriction of caloric intake (to reduce total body fat), and the admini stration 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 br eakdown 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 treatment s that were effective in enhancing the biliary and intestinal elimination of PBB residues (Polin et al. 19 91; Rozman et al. 1982). Polin et al. (1991) found that dietary intervention to reduce PBBs was dose de pendent; treatment with 10% mineral oil and a 45%

reduction in food intake resulted in a 69 and 23% redu ction in body burden in rats fed PBBs at dietary concentrations of 0.1 and 100 ppm, respectively (Pol in 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 treat ment alone had no effect in reducing PBB body burden. A 3-week treatment regimen that included di etary supplements of polyunsaturated oil, vitamins, and minerals, and heat stress has been applied in a pi lot study to seven human subjects that were known to have been exposed to PBBs; following treatment, stat istically 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 stud y (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 fo r 16 individuals who developed symptoms following exposure to PCBs and polychlorin ated dibenzofurans (Imamura a nd 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 info rmation for PCBs, mobilization of PBBs or PBDEs from adipose tissue is not recommended in indivi duals 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 m echanism 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

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that interfering with the initial step, binding to the r eceptor, 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 suc cessful 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-me diated effects of PBBs after exposure has occurred is largely untested. Fo r PBDEs, the ether bridge precludes congeners from readily adopting the coplanar (dioxin-like) conforma tion 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 anta gonists 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 oxid es (Kohli and Safe 1976; Kohli et al. 1978) that may alkylate critical cellular macromolecules and result in injury. PBDE-induced decreases in thyroid T 4 hormone, which can affect neurobehavioral developmen t, are likely to involve multiple mechanisms (see Section 5.5.3). These include induc tion of hepatic microsomal enzymes, particularly UDPGT, which can

increase the rate of T 4 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 T 4 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 im pact of hydroxy-metabolites on serum T 4 needs further

clarification. Effects of PBDEs on the function a nd development of the nervous system could also involve disruption of calcium homeostatic mechanis ms 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 s ynaptosomes (Mariussen and Fonnum 2002, 2003; Mariussen et al. 2003a) and cholinergic functions (Vib erg et al. 2001b, 2002a, 2002b), and/or free radicalinduced neuronal death (Reistad et al. 2002). Clini cal 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 whetherPBBs and PBDEs 317

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adequate information on the health effects of PBBs a nd PBDEs is available. Where adequate information is not available, ATSDR, in conjunction with th e 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 substan ce-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 f illed. In the future, the id entified 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, or al, 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 Relate d to Toxicological Profiles (Agency for Toxic Substances and Disease Registry 1989), is substa nce-specific information necessary to conduct comprehensive public health assessments. Generally , ATSDR defines a data gap more broadly as any substance-specific information mi ssing from the scientific literature.

As indicated in Figure 5-6, information is availa ble regarding systemic, immunological, neurological, developmental, reproductive, and carci nogenic effects of PBBs in humans. The information on effects in humans is derived from the Michigan contamination ep isode that involved chronic-duration oral exposure

to contaminated food and from oc cupational exposure data in whic h it was assumed that exposure was predominantly through skin contact, although in halation exposure cannot be ruled out.

Information on health effects in animals is extensive a nd available for all effect categories, but is nearly completely limited to oral exposure studies, whic h appears to reflect experimental practicality and concern for what is thought to be the most prev alent and likely route of environmental exposure.PBBs and PBDEs 318

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### Figure 5-6. Existing Information on Health Effects of PBBs

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A limited amount of information is available on th e 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 expo sure studies in animals. In general, the health effects of PBDEs are less adequately studied than fo r 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; Co rbett et al. 1978; Fraker 1980; Gupta and Moore 1979; Kimbrough et al. 1981), and thyroid effects occu rred at doses as low as those causing liver effects. Insufficient acute data exist to definitely establish if th e thyroid effects are more critical than effects in the liver, but extensive data on thyroid effects from l onger 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 estab lish the most sensitive target and species for acute exposure. Tests with monkeys, guinea pigs, and mi nk would be informative be cause intermediate- and chronic-duration studies indicate that these speci es are more sensitive than the rat and that endocrinological effects are particularly sensitive end points.

Information on toxic effects of acute-duration exposur e to PBBs by routes other than oral are limited to data on hepatic, renal, dermal, and ocular effects of inhalation and de rmal 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 dela yed lethality. Limitations in the animal database include inadequate reporting (e.g., numbers of animal s 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 lacki ng. 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 iden tifying target organs and examining the dose-responsePBBs and PBDEs 321

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relationship following acute inhalation and dermal expos ure 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 docum ented 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 i ndicate 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 PB DEs that need to be further investigated (see discussions of data needs for Immunotoxicity and Neurot oxicity). Studies in other species would help to clearly establish the most sensitive target and species fo r acute exposure, as well as which animal toxicity

data are the most relevant to humans and useful fo r 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 suffici ently 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 brom inated 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 identifie d in these studies, but MRL estimation is precluded by

inconsistencies between the studies and a lack of information on thyroid horm one levels. Additional dose-response studies are needed to pr ovide 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 th e 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

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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; Lambrech t 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; Wa ritz et al. 1977; Werner and Sleight 1981). Studies have been performed with various sp ecies (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. Hema tologic 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 reproductiv e and developmental effects occurred in monkeys at a lower dosage. The serious nature of the devel opmental toxicity (fetal death) precludes derivation of

an intermediate-duration oral MRL. Additional in termediate-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 interspeci es differences could help to better characterize the relative sensitivity of monkeys and humans, particular ly the possibility that monkeys are more sensitive than humans, as indicated by the high reproductive/deve lopmental toxicity of PBBs in this species that has not been noted in PBB-exposed workers or th e 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; Norri s et al. 1975a; Waritz et al. 1977). Some inhalation data are available for octabromobiphenyl and decab romobiphenyl mixtures and some dermal data are available for octabromobiphenyl mixture, but interm ediate-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. A lthough 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 decabromobipheny l 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 mo re 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 thePBBs and PBDEs 323

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low volatilization potential of PBBs. The intermed iate-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 si tes. Intermediate-dura tion inhalation and dermal exposure studies of PBB-contaminated soil (e.g., bioava ilability studies) that identify thresholds would be especially useful for risk assessme nt 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 an d histological alterations in both organs and changes in serum levels of thyroid hormones, particularly decreases in serum T 4 (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 intermedia te 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 intermed iate 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 eff ects of lower brominated PBDEs have been obtained from oral studies in rats. It is speculated that the ex tent that PBDEs affect circulating levels of thyroid

T 4 or T 3 might vary with species, and rats are often rega rded as more sensitive than humans. As indicated in the Comparative Toxicokinetics section below, spec ific 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-Daw ley 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 n ecropsied. 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 co rpora 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 malfor mations and ossification va riations (approximatelyPBBs and PBDEs 324

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half of each litter), and evaluated for visceral malf ormations (remaining fetuses). No treatment-related effects on any maternal or fetal endpoints were ob served, indicating that 1,000 mg/kg/day was the NOAEL for maternal and developmental toxicity.

One intermediate-duration oral toxicity study has b een conducted for high purity decaBDE. In this study, multiple dose levels of a co mmercial decaBDE product (94

B 97% pure) was fed to rats and mice for 13 weeks (NTP 1986). Comprehensive histological examin ations were performed, but were limited to the control and high-dose groups, and no hematology, clini cal chemistry, or urine indices, or thyroid hormone

levels were evaluated. There were no compound-related clinical si gns, deaths, body weight or food consumption changes, gross pathology, or histopathology, i ndicating 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 th is 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 intermed iate-duration oral MRL for decaBDE.

The inhalation database for intermed iate-duration exposure to PBDE s 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 estimati on of an intermediate-dur ation 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-dur ation 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-contamin ated soil following oral, inhalation, and dermalPBBs and PBDEs 325

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exposure (i.e., bioavailability studies) would be particularly useful. Ev aluations of the thyroid would be particularly informative because intermediate-durati on 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 (94

B 97% pure) was fed to rats and mice for 103 weeks (NTP 1986). Comprehensive gross and histological exam inations 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 throm bosis in the same tissue, precluding estimation of an MRL. Additional chronic dose-response inform ation 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 inci dences 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 lo wer brominated PBDEs is available from a study in which rats were fed a 77.4% pure commercial d ecaBDE 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, hema tology, clinical chemistry, urine indices, and comprehensive histological examinations showed no e xposure-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 cha nges 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 st udies of pentaBDE and/or octaBDE are needed to provide adequate bases for MRL derivation and can cer assessment for lower brominated PBDEs. Evaluations that include the thyroid and neurobehavi oral end points would be particularly informativePBBs and PBDEs 326

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because acute and intermediate-duration oral studies indicate that the thyroi d and developing central nervous system are particularly sensitiv e 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; N TP 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 pot encies 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 examina tion 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) fo r 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 chro mosomal 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 increase d recombinogenic activity in Chinese hamster SPD8 and Sp5V79 cells (Helleday et al. 1999). Althoug h the weight of available evidence indicates that decaBDE is not genotoxic, studies using lower bromin ated mixtures and a wider variety of assay types would help to better characterize the genotoxic potential of PBDEs.PBBs and PBDEs 327

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Reproductive Toxicity.

Polybrominated Biphenyls. A limited amount of information is av ailable regarding reproductive effects in humans after exposure to PBBs. No evidence fo r PBBs-related effects on sperm counts, motility, or sperm morphology was found in a group of male Mich igan workers exposed to PBBs by inhalation or dermal contact (Rosenman et al. 1979). No relations hip was found between serum levels of PBBs and the frequency and duration of lactation in women expo sed 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 Ringe r 1979; Ringer et al. 1981) or in the F 1 or F 2 generations of female F 0 rats fed PBB-containing diets during pos timplantation gestation through weaning (McCormack et al. 1981), implantation was completely blocked in 40–67% of fe male rats exposed by gavage to PBBs between GDs 0 a nd 14 (Beaudoin 1979). Additionally, a lengthening of the menstrual cycle and prolonged implantation bleeding with decr eased 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 reproduc tive effects in monkeys was the lowest tested in any intermediate-duration study of PBBs. In add ition, 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. Hist opathological 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 anim al data suggest that PBBs may cause adverse effects on reproductive organs and their function(s) an d that reproductive organ functions during the early phases of pregnancy may be particularly sensitive to P BBs. Additional studies in animals exposed by oral and other routes, including multi-generation studi es 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 relati onships. Studies elucidating the NOAEL region and relative susceptibility of sensitive species (e.g., m onkeys) to reproductive and developmental effects would be particularly useful, as these data could en able 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 st udies designed to assess effects on fertility in bothPBBs and PBDEs 328

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sexes, would better characterize the reproductive toxic pot ential 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 neur opsychological 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 conc lusively 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; Lu ster et al. 1980), fetal malformations (Beaudoin 1977; Corbett et al. 1975; Waritz et al. 1977), growth re tardation in offspring (Allen et al. 1979; Aulerich and Ringer 1979; Corbett et al. 1975; Groce and Kimb rough 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 th e lowest dosage tested in any study of PBBs, a chronic oral MRL could not be calculated; studi es 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 ornot the developmental toxicity of PBBs is routespecific. Studies on the mechanism(s) of action of PB Bs in different animal species may provide a better understanding of the physiological and biochemical b asis for the developmental toxicity of PBBs and a better basis for extrapolating from animal data in th e 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 ra ts and rabbits, although fetotoxic effects, including skeletal ossification variations at maternally toxi c 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; Norri s et al. 1975a, 1975b; WIL Research LaboratoriesPBBs and PBDEs 329

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1986). Effects of gestational exposure to lower br ominated PBDEs included minimally increased postimplantation loss in rats and increased skeletal vari ations in rabbits at octaBDE doses as low as 10 and 15 mg/kg/day, respectively (Argus Research Labor atories 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 (H ardy 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 c ongeners, 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, 2 001a, 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 reporte d 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 th ese 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, co mpared 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 mit ogenic 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 (V os 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 relativel y low vapor pressure of PBBs, inhalation is not a predominant route of exposure. Additional oral st udies using a battery of immunological tests would be useful to further define the immunological effects of PBBs.PBBs and PBDEs 330

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Polybrominated Diphenyl Ethers. Information regarding the immunosuppressive potential of PBDE mixtures is essentially limited to evidence from acu te-duration oral studies in animals exposed to relatively high doses of pentaBDE. The plaque-formi ng 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 f ound 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 Thuva nder 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 (Darneru d and Thuvander 1998). In a 2-year study, NTP (1986) reported that exposure to 2,240 mg/kg/day of d ecaBDE 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 studi es 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 certa in 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 predomin ately 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 reporte d 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 Ra e 1983; Seagull 1983; Weil et al. 1981). Subtle

effects in neurobehavioral tests were found in rodent s, 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 l earning 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 spin al nerve tissue revealed no abnormalities in rats orPBBs and PBDEs 331

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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 se vere 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 assess ed by motor activity, habituation, and auditory startle response, although suggestive decreases in fear conditioning were observed.

Neurobehavioral effects of individua l 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 experime ntal designs and evaluated spontaneous motor behavior and swim maze pe rformance at 2–6 months of age. The findings collectively indicate that the nervous system is a target of particular PBDE c ongeners during a defined critical phase of neonatal brain development, as shown by mild impairments in spont aneous 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 ge otaxis, 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 a ltered spontaneous behavior.PBBs and PBDEs 332

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Additional oral studies using comprehensive neurobe havioral test batteries are needed to better characterize the potential for PBDEs to cause neur otoxic effects in humans. Based on the limited available information (predominantly study abstract s), studies evaluating effects of exposure to lowerbrominated PBDEs and decaBDE commercial mixtur es 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 contaminat ion episode or who were exposed occupationally in the manufacture or distribution of PBBs have not pr ovided conclusive evidence that detectable effects have occurred as a result of exposure to PBBs (Ande rson et al. 1978c, 1979; Barr 1980; Henderson et al. 1995; Hoque et al. 1998; Humble and Speizer 1984; Landr igan 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 contaminati on episode (Thomas et al. 2001). A relationshipbetween increasing serum levels of PBBs and increasing risk of breast cancer was indicated in casecontrol studies of women exposed during the Michig an episode (Henderson et al. 1995; Hoque et al. 1998), but the results are only suggestive due to f actors 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 asso ciation 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 wast e site, the nearby population should be studied for both exposure and effect data.

Polybrominated Diphenyl Ethers. A limited amount of epidemiologi cal information is available for

PBDEs. Plasma levels of vari ous organohalogen compounds, including the congener 2,2',4,4-tetraBDE, as well as serum hormone levels (free and total T 3 and T 4 , TSH, free testosterone, follicle-stimulating hormone, lutenizing hormone, and prolactin), were an alyzed in 110 men who consumed fatty fish from the Baltic Sea (Hagmar et al. 2001). There was a w eak negative correlation between 2,2',4,4'-tetraBDEPBBs and PBDEs 333

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(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-contro l study of 77 Swedish men and women (Hardell et al. 1998; Lindstrom et al. 1998). 2,2'4,4'-TetraBDE was u sed as a marker for total PBDE exposure in both of these studies. No epidemiological data on PBDE s 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 carcinogeni c 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 effect s as well as the relationship between PDBE body burden and the observed effects. Considering the possi bility 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 e ffects 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 PB B 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). Furt her 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 w ould provide valuable information that could lead to early detection of PBB exposure.PBBs and PBDEs 334

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A potential biomarker of exposure to PBBs is related to their effect on the thyroid gland. Effects in exposed workers included increased serum th yrotropin, low or borderline low serum T 4 , and increased thyroid antimicrosomal antibody titers (Bahn et al. 19 80), and effects in exposed rats included reduced

levels of serum T 4 and T 3 (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 a nd Sleight 1981). Additio nal studies could better characterize thyroid effects of PBBs in humans a nd develop specific correlati ons between levels and duration of exposure and alterati ons in serum hormone levels.

Polybrominated Diphenyl Ethers. PBDEs also accumulate in adipose tissue, serum, and breast milk of the general population due to their lip ophilic 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 McDona ld 2000). Breast milk monitoring programs are needed to provide timetrend data and to verify findings that PBDE levels have been e xponentially 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 u seful 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 so luble 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 T 4 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 biom arker may not be reliable, as thyroid changes are not specific to exposure to PBDEs and the effects ass ociated 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 human s and develop specific correlations between levels and duration of exposure and alte rations in serum levels of T 4 .

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 healthPBBs and PBDEs 335

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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. Cont inued follow-up studies on the Michigan cohort would provide information on effects that may have a long la tency, 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 T 4 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. Leve ls of CYP enzymes and thyroid hormones, however, are not specific for PBB exposure. Further studies designe d to identify specific biomarkers of effects of PBBs would facilitate medical surveillance leading to ear ly detection of potentially adverse health effects and possible treatment. Congener-specific analysis ma y 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 T 4 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 T 4 could be considered to indicate potential impairment of health.

Absorption, Distribution, Metabolism, and Excretion.

Polybrominated Biphenyls. There are no quantitative data regard ing absorption in humans via the inhalation route, but data from o ccupationally exposed individuals a nd individuals who ingested food contaminated with PBBs suggest that exposure by th e 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 ofa commercial PBB mixture (2,2,'4,4',5,5'-hexabromobiphenyl) is efficiently absorbed by the or al route (Matthews et al. 1977; Tuey and Matthews 1980). Data regarding absorption after inhalation expo sure 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 administ ered by the inhalation, oral, and dermal routes, and for various exposure periods. Such studies coul d 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

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In addition, studies with different PBB mixtures coul d help determine possible interaction effects among congeners that could affect absorption.

Distribution data are limited to qualitative informatio n derived from cases of accide ntal ingestion of food contaminated with PBBs, cases of occupational e xposure 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 regard ing 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 we re 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 metabo lic fate of PBBs. There are studies regarding the metabolism of some PBB congeners after oral administ ration to rats (Sparling et al. 1980), rabbits (Kohli et al. 1978), and pigs (Kohli and Safe 1976). Ho wever, 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 dis position of highly brominated congeners. A limited

amount of information is available on the metabolis m of PBBs in farm animals (e.g., dairy cows, chickens). This is a data gap because people expo sed 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 (Brillia nt et al. 1978; Eyster et al. 1983; Jacobson et al. 1984; Landrigan et al. 1979). Fecal excretion of un absorbed 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

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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 de rmal exposure were not located, there is no reason to

believe that results from additional studies w ould reveal different patterns of excretion.

Polybrominated Diphenyl Ethers. A limited amount of data is availa ble on the toxicokinetics of PBDEs. There are data gaps in a number of areas, partic ularly for octaBDE and pentaBDE commercial mixtures and the tetra and hexa congeners that are most pr evalent in the environmen t. Quantitative absorption studies could corroborate the conclusions on oral upta ke 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 activ e hydroxylated BDEs. There are still data gaps in the toxicokinetics of decaBDE, including an incomp lete understanding of the debromination of decaBDE

to lower brominated BDEs.

Comparative Toxicokinetics.

Polybrominated Biphenyls. The data suggest that there are qualit ative 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 hi ghly 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 patte rn of bromination (Kohli et al. 1978; Matthews et al. 1977; Sparling et al. 1980). Analysis of the excreta of humans e xposed in the workplace and near hazardous waste sites would provide information regarding biotransformation and elimination ki netics in humans. In addition, similar target organs have been identified across an imal 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 huma ns (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 compone nt of FireMaster mixtures, 2,2',4,4',5,5'-hexabromobiphenyl (BB 153), in the rat (Tuey and Matthew s 1980). The serum mean half-life of 6.5 yearsPBBs and PBDEs 338

5. HEALTH EFFECTS

predicted using this model is shor ter 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 di fferences 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 extrapola ting 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 depende nt on the specific congener or mixture studied, and pharmacokinetic modeling studies could he lp 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 T 4 or T 3 might vary with species and rats are often regarded as more sensitive than humans. Th e main basis for this opinion seems to be studies showing that PBDEs affect binding of thyroid hormon es to transthyretin (TTR), the primary transport protein in rats. Because TTR is not the major tran sport 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 ra ts is thought to be rela ted 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 horm one 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 me thods 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-e xposed humans has not been fully researched.PBBs and PBDEs 339

5. HEALTH EFFECTS

The mechanism of toxic action of PBBs is not comple tely understood and no methods exist to block the toxic response due to exposure to PBBs. A more comp lete characterization of the cytosolic AhR protein, to which some PBB congeners are thought to bind, a nd understanding of physiological effects of receptor blockage would be useful for the possible identifica tion 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 s ubjectively 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 Mi chigan 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 thyroi d gland is an unequivocal target of PBBs in adult animals. No information was located on possible imm unological effects of PBBs in children, and data in adult humans are limited and largely inconclusive (Bek esi et al. 1978; Landrigan et al. 1979; Roboz et al. 1985; Silva et al. 1979; Stross et al. 1981), but expos ure to PBBs caused altered immune responses in a variety of animal species, which suggests that child ren may also be affected. Continued assessment of children exposed to PBBs during the Michigan c ontamination episode, with particular emphasis onPBBs and PBDEs 340

5. HEALTH EFFECTS

evaluation of cognitive abilities, thyroid function, an d 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 altera tions have been observed in animals following preand/or postnatal exposure to co mmercial PBDE mixtures and single PBDE congeners, indicating that these are potential effects of concern in expo sed children. Serum levels of thyroid T 4 and T 3 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). A lterations 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) (Erikss on 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 th e potential susceptibility of children to the effects of PBDEs on the thyroid and neurodevelopm ent, particularly considering th e 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 anim als 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 effect s of PBBs and PBDEs, as identified in the Federal Research in Progress database (FEDRIP 2002) and the we bsites of various U.S. government agencies, are

listed in Table 5-9.

The Great Lakes Chemical Corporation is collaborativ ely working with Health Canada to assess potential effects of pentaBDE using a three-tier study appro ach (Biesemeier 2004). The first tier study, a 28-day repeated oral (gavage) toxicity st udy 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 secondPBBs and PBDEs 341

5. HEALTH EFFECTS



### Table 5-9. Ongoing Studi es 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/EPA a 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/EPA a 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 2004dPBBs and PBDEs 342

5. HEALTH EFFECTS



### Table 5-9. Ongoing Studi es on the Health Effects of PBBs and PBDEs

Investigator Affiliation Research description Sponsor Source Robertson L Universi ty of Kentucky Medical Center, Lexington, Kentucky Laboratory synthesis of pure PBDE and PCDE congeners and their metabolites NIEHS/EPA a FEDRIP 2002

Sikka HC State University of New York College at Buffalo, Buffalo, New York Disposition and metabolism of PBDEs in fish NOAA b FEDRIP 2002

Trosko JE Michigan State University, East Lansing, Michigan Epigenic effects of PBBs and other environmental toxicants on cellular communication pathways NIEHS/EPA a 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 xenobiotic s 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. En vironmental Protection Agency; EU = European Union; NIEHS = National Institute of Environmental Health Sciences; NOAA = National Oceanic and Atmospheric Administration; PCB = polyc hlorinated biphenyls; PBPK = physiologically based pharmacokinetic; USDA = U.S. Department of AgriculturePBBs 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 prot ocol is being drafted with a pha rmacokinetic 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 th e third tier study, a two-ge neration reproductive study in rats with endocrine, developmen tal neurobehavioral, and other end poi nt 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 in cluded when describing PBBs. The general chemical

structure of PBBs is shown below:



### 1' 2' 3'

4'

5' 6' 1 2 3

4

5 6 Br n Br m where m + n = 1–10

It can be seen from the structure that a large num ber 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 th e number of bromine substituents, there are 10 homologous groups of PBBs (monobrominated th rough 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 fo r 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 de termined 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 larg er 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 di hedral angle is small, but >0°) or “near” planar configuration. ThesePBBs and PBDEs 346

6. CHEMICAL AND PHYSICAL INFORMATION

molecules are referred to as planar or coplanar congeners. The ben zene 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 arrange d 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 wi dely 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 show n 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 retard ants using different numerical designations. Other former producers of PBBs in the United States in cluded White Chemical Corporation (Bayonne, New Jersey) and Hexcel Corporation (Sayreville, New Jers ey), 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 Fire Master 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 of ten included when describing PBDEs. The general

chemical structure of PBDEs is shown below:

O

Br n Br m 

### 1' 2'

3'

4'

5' 6' 1 2

3

4 5 6

where m + n = 1–10PBBs 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’,4PBBs 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’,5PBBs 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’,6PBBs 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’,6PBBs 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 C AS numbers were identified for the PBDE class.

c WHO 1994b d WHO 1994aPBBs and PBDEs 353

6. CHEMICAL AND PHYSICAL INFORMATION



### Table 6-2. Chemical Id entity of Selected PBBs a

Characteristic Hexabromobiphenyl Octabromobiphenyl Decabromobiphenyl 2,2',4,4'5,5'-Hexabromobiphenyl Synonym(s) FireMaster BP-6 b ; FireMaster FF-1 b Bromkal 80 b Flammex B 10 b ; Adine 0102 b ; Berkflam B 10 b 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 C 12 H 4 Br 6 C 12 H 2 Br 8 C 12 Br 10 C 12 H 4 Br 6 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 thei r 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 = Nationa l Cancer Institute; NIOS H = National Institute for Occupational Safety and Health; OHM/TADS = Oil and Ha zardous Materials/Technical Assistance Data System; RTECS = Registry of Toxic Effects of Chemical SubstancesPBBs and PBDEs 354

6. CHEMICAL AND PHYSICAL INFORMATION

It can be seen from the structure that a large num ber of brominated compounds are possible. The 209 possible compounds for PBDEs are called “congener s”. However, the number of PBDE congeners that actually exist in commercial PBDE mixtures ar e much less compared to PCBs. Typically, only a subset of the 209 possible congeners is observed for PB DEs. 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-conge ners 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 repr esentative 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-s ubstituted for a diphenyl ether molecule to assume a planar or near planar configuration. Ha logen substitution of the diphenyl ether molecule in the ortho position (2,2’,6,6’) will force the aromatic rings orthogonal to one anot her (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 conge ners (Hardy 2002).

Like PCBs, the 209 congeners for PBDEs are arra nged 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 number s assigned by Ballschmiter and Zell (which are also referred to as congener, IUPAC, or BZ numbers) are wi dely 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 tradePBBs and PBDEs 355

6. CHEMICAL AND PHYSICAL INFORMATION

names used by producers from Europe and Japan fo r the BDE mixtures. The chemical identities of commercial mixtures of penta-, octa-, and decabro modiphenyl 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 = Polybr omodiphenyl ethers = PBDEs or PBDPEs polybrominated diphenyl oxides = Polybr omodiphenyl oxides = PBDOs or PBDPOs

For consistency in this document, pol ybrominated 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 ch emical 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, aPBBs and PBDEs 356

6. CHEMICAL AND PHYSICAL INFORMATION



### Table 6-3. Chemical Iden tity of Technical PBDEs

Characteristic Pentabromodiphenyl ether Octabromodiphenyl ether Decabromodiphenyl ether Synonym(s) Pentabromodiphenylether; pentabromodiphenyl oxide; pentabromobiphenyl oxide; benzene, 1,1-oxybis, pentabromo derivative Octabromodiphenylether; 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 C 12 H 10-y Br y O where y=4–6 C 12 H 10-y Br y O where y=6–9 C 12 Br 10 O

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 = Nationa l Cancer Institute; NIOS H = National Institute for Occupational Safety and Health; OHM/TADS = Oil and Ha zardous Materials/Technical Assistance Data System; PBB = polybrominated biphenyl; RTECS = Registry of Toxic Effects of Chemical SubstancesPBBs and PBDEs 357

6. CHEMICAL AND PHYSICAL INFORMATION



### Table 6-4. Physical and Chemi cal Properties of Selected PBBs a

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 °C b (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 μ g/L c 20–30 μg/L Insoluble 11 μg/L d 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 K ow 6.39 e 5.53 8.58 f 9.10 (estimated) d Log K oc 3.33–3.87 g No data No data 5.088 d Vapor pressure 5.2x10 -8 mmHg at 25 °C h ; 5.6x10 -6 mmHg (liquid subcooled) c 7x10 -11 mmHg at 28 °C i No data 7.6x10 -5 mm Hg at 90 °C d

Henry's law constant 3.9x10 -6 atm-m 3 /mol j ; 1.38x10 -6 atm-m 3 /mol c No data No data 5.7x10 -3 atm-m 3 /mol d

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'-hexabro mobiphenyl are given as 7.20 (Chessells et al. 1992) and 8.09 (Anliker et al. 1988), respectively. g Estimated from the Freundlich adsorption c onstants 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 Fi reMaster 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 (S undstrom 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 a nd FireMaster BP-6 is that they contain >50% of the congener BB 153. The second most abundant cong ener is 2,2',3,4,4',5,5'-heptabromobiphenyl (BB 180). A detailed analysis of FireMaster BP-6 (l ot 7062) was able to separate 22 congeners of PBBs that included four tri, five tetra, three penta, seven hexa, and three hepta congene rs 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 th e 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 hexa bromonaphthalene (Di Carlo et al. 1978); however, at a detection limit of 0.5 ppm, brominated dioxins and dibenzofurans we re 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, a nd 5.7–6.0% decabromobiphenyl (Norris et al. 1973; Waritz et al. 1977). Notably, the major component of commercial octabromobiphenyl was nonabromobiphenyl, and not oc tabromobiphenyl. 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 ra nge of 600–900 °C in the absence of oxygen produced bromobenzenes and brominated biphenyls as key produc ts, 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-tetrabro modibenzofuran 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 ratePBBs and PBDEs 359

6. CHEMICAL AND PHYSICAL INFORMATION



### Table 6-5. Identified P BB 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 1988PBBs and PBDEs 360

6. CHEMICAL AND PHYSICAL INFORMATION



### Table 6-5. Identified P BB 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 1980PBBs and PBDEs 361

6. CHEMICAL AND PHYSICAL INFORMATION

of PBB hydrolysis under much milder environmen tal conditions remains unknown (Pomerantz et al. 1978).

Hexabromonapthalene has been identified as a toxi c contaminant of Firemaster BP-6 or FF-1 at concentration levels of approximate ly 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-hexa bromonaphthalene.

Polybrominated Diphenyl Ethers . Information found in the litera ture regarding the physical and chemical properties of selected technical PBDE mixt ures is presented in Table 6-6. Recent information regarding the vapor pressure, water solub ility, Henry’s Law c onstant, and log K ow 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 comme rcial mixture pentabromodiphenyl ether denotes the main component of the mixture contains the pe ntabromodiphenyl ether homolog. However, the commercial pentaBDE mixture actually contai ns tetrabromodiphenyl ether (24–38%) and pentabromodiphenyl ether (50–62%) homologs with small amounts of hexabr omodiphenyl ether (4–8%) and trace amounts of tribromodiphenyl ether (0–1%) homologs. In this document, the commercial mixture of pentaBDE may be called “the commercia l 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-, he pta-, 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 diphen yl 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 rema inder 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 octaBD E (e.g., DE-79) is <10% of nonaBDE, <33% of octaBDE, <45% of heptaBDE, and <12% of hexa BDE (i.e., 2% of BDE 154 and 14% of BDE 153)PBBs and PBDEs 362

6. CHEMICAL AND PHYSICAL INFORMATION



### Table 6-6. Physical a nd Chemical Properties of Technical PBDE Mixtures

Property Pentabromodiphenyl ether Octabromodiphenyl ether Decabromodiphenyl ether Molecular weight Mixture Mixture 959.22 a Color Clear, amber to pale yellow a Off-white a Off-white a

Physical state Highly viscous liquid Powder Powder a Melting point -7 to -3 °C (commercial) b 85–89 °C (commercial) c ; 200 °C (range, 167– 257) a ; 79–87 °C a ; 170– 220 °C a 290–306 °C a

Boiling point >300 °C (decomposition starts above 200 °C) a,b Decomposes at >330 °C (commercial) c Decomposes at >320, >400, and 425 °C a Density (g/mL) 2.28 at 25 °C a ; 2.25–2.28 b 2.76 a ; 2.8 (commercial) c 3.0 a ; 3.25 a Odor No data Faint a Odorless a 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 μ g/L (heptabromodiphenyl ether component) c <0.1 μ g/L g

Organic solvent(s) 10 g/kg methanol; miscible in toluene d acetone (20 g/L); benzene (200 g/L); methanol (2 g/L) all at 25 °C a acetone (0.05%), benzene (0.48%), methylene bromide (0.42%), xylene (0.87%), and toluene (0.2%) e Partition coefficients: Log K ow 6.64–6.97 d ; 6.57 (commercial) b 6.29 (commercial) c 6.265 e

Log K oc 4.89-5.10 f 5.92-6.22 f 6.80 f Vapor pressure 2.2x10 -7 –5.5x10 -7 mm Hg at 25 °C d ; 3.5x10 -7 mm Hg (commercial) b 9.0x10 -10 –1.7x10 -9 mm Hg at 25 °C d ; 4.9x10 -8 mm Hg at 21 °C (commercial) c 3.2x10 -8 mm Hg g ; 3.47x10 -8 mm Hg e

Henry’s Law constant (atm-m 3 /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 °C b,d Decomposes above 330 °C (commercial) c Not applicable a

Flashpoint No data No data None Flammability limits Not applicable (flame retardant) b,d Not applicable (flame retardant) c Non-flammable aPBBs and PBDEs 363

6. CHEMICAL AND PHYSICAL INFORMATION



### Table 6-6. Physical a nd Chemical Properties of Technical PBDE Mixtures

Property Pentabromodiphenyl ether Octabromodiphenyl ether Decabromodiphenyl ether Conversion factors 1 ppm=23.48 mg/m 3 at 20 °C d No data No data

Explosive limits None b,g None c 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 Congeners a,b

Congener Vapor pressure (mm Hg) a,c Water solubility ( μ g/L) a Henry’s Law constant (atm m 3 /mol) a Log K ow b

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 etherPBBs 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 pentaB DE (i.e., 8% of BDE 100, 43% of BDE 99), 24– 38% of tetraBDE (i.e., <1% of BDE 66, <1% of BD E 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 gi ven 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-substitu ted brominate dibenzop -dioxins and dibenzofurans revealed no detectab le 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 samp les, there are not measurable quantities of PBDDs/PBDFs, there are some materials that have re ported quantifiable levels of these contaminants. For example, hexabromodibenzofurans have been detected in commercial decaBDE mixtures at concentrations as high as 200 μ g/kg. In other PBDE mixtures (e.g., tetra- to hexaBDEs), the sum of tetra-, penta-, and hexabromodibenzofurans we re reported at a concentrations of 8,000 μ g/kg. In addition, tetra- and pentabromo- p -dibenzodioxins have been measured in commercial decaBDE at concentrations

of 0.05 and 0.35 μ g/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 concentr ations can be generated during pyrolysis of decabromodiphenyl ether (decaBDE) in the temperatur e 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 me thods that could not differentiate between PBDD/Fs formed (e.g., 2,3,7,8-substituted c ongeners) 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 Comme rcial 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%) 0–1% – – – – TriBDE (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 detectedPBBs and PBDEs 367



### 7. PRODUCTION, IMPORT/E XPORT, 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 (S undstrom et al. 1976a). In one process, biphenyl

is brominated with 0–20% stoichiometric excess of br omine chloride (e.g., slightly more than 10 mol of bromine chloride may be reacted with 1 mol of bi phenyl to obtain decabromobiphenyl) in the presence of iron or a Friedel-Crafts catalyst (e.g., aluminum chlo ride). In another process, biphenyl is dissolved in ethylene bromide solvent and reacted with bromin e in the presence of a catalyst (either aluminum chloride or bromide) (Neufeld et al. 1977). Resear ch 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; Robert son et al. 1983b). Methods for laboratory scale synthesis of 42 congeners of brominated bipheny ls are also available (Sundstrom et al. 1976b).

The commercial production of PBBs began in 1970. Approximately 13.3 milli on pounds of PBBs were produced in the United States from 1970 to 1976. Only three commercial PBB products were manufactured (i.e., hexabromobiphenyl, octabromobi phenyl, and decabromobiphenyl) and these three products were based on a limited number of congene rs (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 d ecabromobiphenyl in the United States until 1979 (IARC 1986; Neufeld et al. 1977). Shortly after th e 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 Stat es (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 Ge rmany until mid-1985. Until the year 2000, the onlyPBBs and PBDEs 368

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

PBB in commercial production was decabromobiphe nyl, 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 a nd 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 vol ume, 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 octabromodipheny l 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 pr oduced 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 commercia lly in the United States as DE-83R and Saytex

102E. Dead Sea Bromines/Eurobrome and Tosoh are cu rrently producing PBDEs in the Netherlands and Japan, respectively (WHO 1994a). No current estim ates 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 a nd 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 ofPBBs and PBDEs 369

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



### Table 7-1. Total Market Demand of PBDEs by Regi ons of the World in 2001 a

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 mi xture; octaBDE = commercia l octabrominated diphenyl ether mixture; pentaBDE = commercial p entabrominated diphenyl ether mixturePBBs and PBDEs 370

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

pentaDBEs and octaDBEs by the end of 2004. Sin ce 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 ma nufacture 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 Tabl e 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 manufactu re 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, TR I data have certain limitations. TRI data reflect releases and other waste management of chemicals, a nd not exposures of the public to those chemicals. TRI data alone are not sufficient to determine expos ure 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 we re 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 St ates. In the past, PBBs were used as additive flame retardants to suppress or delay combus tion. Additive flame retardants are added to the polymer material, but are not chemically incorporated into the polymer matrix. Because PBBs are notPBBs and PBDEs 371

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



### Table 7-2. Facilities th at Produce, Process, or Use Decabromodiphenyl Ether

State a Number of facilities Minimum amount on site in pounds b Maximum amount on site in pounds b Activities and uses c 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 ImpurityPBBs and PBDEs 372

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

chemically bound to the polymer matrix, they ma y 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 ma inly in thermoplastics for constructing business machine housings and in industrial (e.g., motor hous ing), and electrical (e.g., radio and TV parts) products. Smaller amounts were used as a fire retard ant 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 th e hexabromobiphenyl was l ess 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 w ith 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 oc toBDE by mass can be sold in the European Union (EU 2003b).

The commercial pentaBDE product is used predomin antly (95–98%) for flame retardant purposes as an additive in consumer products manufactured by the fu rniture industry (ENVIRON 2003a). It is used almost exclusively for flame retard flexible polyuret hane foam (FPUF), which is used in bed mattresses and cushioning in upholstered products. The comme rcial 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 approxima tely 2–3% flame retardant mixture and cushion FPUF contains 3–5% flame retardant mixture (ENVI RON 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 fo und 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 specialtyPBBs and PBDEs 373

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

textiles, printed circuit board components, hydraulic and oilfield completion flui ds, 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 pentaB DE. However, this use was discontinued in the early 1990s. Prior to approximately 1990, the co mmercial 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 carpe ts (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 fi nd its way into the United States (EU 2001). The commercial pentaBDE product is not used in acryl onitrile-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, ri gid polyurethane, rigid polyurethane foams, epoxies, laminates, adhesives, coatings, un saturated polyesters, and flexible polyvinyl chloride (PVC) compounds (Great Lakes Chemical Corporation 2004a).

The commercial octaBDE is used by the plastics indus try as an additive flame retardant for manufactured products. It is used almost exclusively to flame re tard 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, simila r 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 (P BT), and polyamide polymers (EU 2003a). Other possible applications of octaBDE include use as a dditive 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 fl ame retardant used in a variety of polymer applications. Industry information indicates that d ecaBDE is used at loadings of 10–15% weight in polymers and is always used in conjunction with an timony 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 polymer s with end-uses in electrical and electronic equipment (e.g., computers, connectors, electrical boxes, wire and cable, etc.). Examples include polypropylene (forPBBs 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 sy stems, and backcoatings for fabrics (Great Lakes Chemical Corporation 2004c). Upholstered furnitu re in commercial settings in the United States is required to meet federal flammability standards and ma y utilize upholstery textiles that are flame retarded with a backcoating containing decaBDE at 5 mg/m 2 (BFRIP 2002).

Currently, technical decaBDE is the most widely us ed 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 domes tic furniture, and in foam-based packaging (EU 2001). Technical octaBDE is mainly used in th e preparation of acrylonitrile-butadiene-styrene terpolymer (ABS), which is used in the manufact ure of computer and business equipment housings.

Technical octaBDE is also used in adhesives a nd coatings applications (WHO 1994a). Technical decaBDE is primarily used in combination with an timony trioxide in high impact polystyrene (HIPS) applications, such as electronic enclosures (e.g., te levision cabinets). Other minor uses of technical decaBDE are in textile applications to flame retard upholstery fabric (e.g., pol yester fiber additives and coatings for automobile fabric, ta rpaulins, 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 sewer s for every 1,000,000 pounds of PBBs produced at manufacturing sites (Neufeld et al. 1977). The Mich igan 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 generate d 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 tota l 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 la ndfilled at other locations. Contaminated animalPBBs and PBDEs 375

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

carcasses, poultry and eggs, animal feed, butter, ch eese, 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 cessati on of production, all of these products, such as TV cabinet and business machine housings, must have b een disposed of by land filling or incineration (Neufeld et al. 1977). The formation of polybromi nated dioxins (PBDDs) and polybrominated dibenzofurans (PBDFs) during the incineration of plastics cont aining 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 (landf ill), 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 consum er products disposed of by each method of waste

transfer.

Landfill disposal of plastic consum ables containing pentaBDE (e.g., pol yurethane 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 a ll 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 meas ures would also be effective in minimizing the leaching of any PBDE particulates present (EU 2002).

PBDEs in sewage are disposed of in publicly owne d treatment works (POTWs), as indicated by analysis of sewage sludge from various countries. Tetra- a nd penta-BDEs have been detected in sewage sludge from POTWs in the United States with the compositi on 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 sew age 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 sour ce. 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 incinerati on 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/PBDD s in effluents from a municipal incineration plant burning domestic waste materials. Flue g ases, fly ash, and bottom ash reportedly contained PBDFs/PBDDs at concentration ranges of 0.28–3.3 ng/N m 3 , 0.082–13 ng/g, and 0.0058–27 ng/g, respectively. However, modern, properly opera ted 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 rec overy (incineration), industrial treatment works, or POTWs. The types of waste transfer may be diffe rent 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., tho se 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 wa stes may contain toluene from the production process (EPA 1995). Plastic processors typically transfer mo st waste to disposal (landfill), recycling, energy recovery (incineration), and industrial treatment wo rks, while minimal releas es are to POTWs. In contrast, textile processors typically transfer most wa ste to POTWs. This difference in waste transfers between the plastic and textile sectors is because tex tile processors use water in their processing operation and other processors (e.g., pro cessors 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 [Sb 2 O 3 ] and decaBDE) and repeatedly ground and injected molded (e.g., “recycled”) HIPS/decaBDE/Sb 2 O 3 resins both met the requirements of the German Chemicals Banning Or dinance with respect to 2,3,7,8-substituted PBDD/PCDF congeners. These resins were at least 1 order of magnitude below the regulated limit valuesPBBs and PBDEs 377

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

for PBDDs/PCDFs (1 ppb for the sum of four cong eners, 5 ppb for the sum of all eight regulated congeners).PBBs and PBDEs 379



### 8. POTENTIAL FO R 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 with in 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 enviro nmental 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'-hexabrom obiphenyl 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 leac hing of commercial mixtures of PBBs from soil is slow. Leaching studies with four Michigan soils mi xed 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 an nual rainfall in Michigan (Filonow et al. 1976). The

PBBs in commercial mixtures resist both chemical a nd 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 SitesPBBs and PBDEs 381

8. POTENTIAL FOR HUMAN EXPOSURE

Kawasaki 1980; Shelton and Tiedje 1981), although bi otic 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 s ubstitutions bioconcentrate in aquatic organisms such as fish, but the octabromo- and decabromobiphenyls do not bioconcentr ate 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 so il 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 pr edators resulting from such consumption remains

unknown.

PBBs were detected in air, water, sediment, and so il 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 rega rding 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 pop ulation is expected to be zero or very small. Populations near the contaminated farms in Lowe r Michigan may still have low exposures from air, water, and food. The level of PBBs in human ti ssue 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 hu man 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 conge ners (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 ra nge of products (WHO 1994a). PBDEs containing wastePBBs 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 par ticulates (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 te traBDE), which may exist partially in the vapor phase, have the potential for long-range transport in th e atmosphere. The detection of lower brominated PBDEs (e.g., 2,2’,4,4’-tetraBDE [BDE 47]) in remo te regions of the world suggests that long-range transport of these congeners is occurring. Highe r BDE homologs (e.g., decaBDE) will primarily exist near point sources. Biodegradation will not be sign ificant 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 degrada tion 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-pentabromodipheny l ether (BDE 99), are transported globally. Atmospheric, water, and biota levels of PBDEs also tend to be dominated by lo wer 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 a ppear to be increasing in certain areas in Canada and the United States, although data are too sparse to ma ke 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 brom inated 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 lo wer 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 contam inated food. Levels of PBDEs in body tissues and fluids from individuals living in the United States ha ve 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 genera l, decaBDE is not detectable in body fluids and tissues. Occupational exposure to PBDEs occurs pr imarily 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 c ongeners, 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 co mmercial 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 constitu ted about 24,500 metric tons (74%), while technical mixtures of octa- and pentabromodiphenyl et hers were 1,500 and 7,100 metric tons (4 and 22% of this total), respectively (BSEF 2003). PBDEs ma y be released into the environment from their manufacture and use in a wide range of consumer products (WHO 1994a). PBDEs are used as additivePBBs 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 possibilit y that some PBDEs congeners (e.g., lower brominated congeners) may diffuse out of the treated material s to some extent (EU 2001). Waste from products containing PBDEs may be either incinerated as municipa l 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 ex ample, 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 releas es of the commercial pentaBDE product from its production and use in the United States. The commerc ial pentaBDE product is used predominantly (95– 98%) as an additive flame retardant in flexible pol yurethane 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 additiv es (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). Sin ce 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 majo r use of pentaBDE is as a component of polyurethane foam for furniture/seating/automobile use, the poten tial 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 fo am has an open cell structure, which presents a large surface area to the environment and therefore, the potentia l 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 re leased could move into the environment and disperse its components (i.e., commercial pentaBDE mixture) (B FRIP 2002). These form particles are released toPBBs 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 with in 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 meas ures 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 monito rs (ENVIRON 2003b). No quantitative information was available on emissions of octaBDE from production oper ations. The major sources of air emissions are thought to be as a result of grinding and bagging ope rations. 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/inc ineration. 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 duri ng uses appears to be small. Waste remaining in the environment can be considered to be particles (or dust) of polymer produc t, or dust generated from polymer products, that contain octaBDE. These partic les are primarily released to urban/industrial soil, but may also end up in sediment or air. Plastics cont aining 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, octa BDE is not expected to leach to a significant extent from polymers, unless the polymer itself undergoes some form of degradati on, 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 reta rdant 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 qua ntitative 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 substan ce 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) wasPBBs and PBDEs 386

8. POTENTIAL FOR HUMAN EXPOSURE

undertaken. The lack of increase in bromine concentra tion 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 d ecaBDE 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 u nder 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 TR I, 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 treat ment 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 y ear, 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 rele ases 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 Louisi ana. On-site air and wa ter releases make up only

a small fraction of total on-site releases. For exam ple, on-site air emissions in 2000 were ~6.6% of the total; on-site surface water discharges also make up on ly a small fraction of the total (~0.2% of the 2000 total) and are nearly all associated with opera tions 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 Ca rolina, 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 ai r 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 brom inated 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 19 87). 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, collect ed 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 par ticles 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 bePBBs 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 dete rmined 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/m 3 , 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 opera tions. However, the emission to air of vapors from production of octaBDE technical mixtures may be cons idered to be negligible (EU 2003a). OctaBDE technical mixtures are used almost exclusively to fl ame retard ABS terpolymers used in computer casings and monitors (ENVIRON 2003b). No quantitative info rmation was available on emissions of octaBDE technical mixtures to the atmosphere from ABS appli cations. 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 μ m) have been estimated as 0.21% during handli ng 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 compoundi ng stage are also susceptible to dust generation. However, losses at this stage are thought to be lo wer 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 ava ilable 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 re tardants (Zweidinger et al. 1979); the concentration of decaBDE ranged from not detected to 72 ng/m 3 . 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 grin ding 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 theref ore, these losses will be mainly to solid waste (EUPBBs and PBDEs 389

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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 expect ed to settle within the compounding facility and will ultimately be to solid waste or waste water. In add ition to particulates, emissions to the atmosphere due to the volatility of the flame retardant at the elevat ed processing temperatures may be possible. However, due to the high vapor pressure of decaBDE (see Table 6-6), volatilization is exp ected to be negligible.

The commercial decaBDE product is an additive flame reta rdant 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 environm ent 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 dus t. 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 particul ar, 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 incinera tion. 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 μ g/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 collect ed 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 fr om Facilities that Produce, Process, or Use Decabromodiphenyl Ether a

Reported amounts released in pounds per year b

State c Number of facilities Air d Water Underground injection Land Total on-site release e Total offsite release f 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 0PBBs and PBDEs 391

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### Table 8-1. Releases to the Environment fr om Facilities that Produce, Process, or Use Decabromodiphenyl Ether a

Reported amounts released in pounds per year b

State c Number of facilities Air d Water Underground injection Land Total on-site release e Total offsite release f 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 ai r, 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

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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 ma nufacture 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 effluent s 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 Carl o 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, Mich igan 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 industri al and urban effluents were located for the United States. Hale et al. (2002) measured the concentra tion of soil and stream sediments collected near a polyurethane manufacturing plant (near the Dan Ri ver, 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 μ g/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 Janss on 1995; Sellström et al. 1998a). Samples were analyzed for tetraBDEs (50 ng/g dry weight) and pe ntaBDEs (sum of three congeners, 2,300 ng/g dry weight). These PBDEs were found in higher concentr ations 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 al ong 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 sedimen ts increased as samples were collected further downstream where additional industries were located (S ellström 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 ( μ g/kg) dry weight; the concentration of BDE 209 ranged from not detected to 16,000 ng/g ( μ g/kg) dry weight. Allchin et al. (1999) surveyed thePBBs and PBDEs 393

8. POTENTIAL FOR HUMAN EXPOSURE

concentrations of PBDEs in sediments from several ri vers 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 refere nce 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 con centrations 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 Sker ne downstream of a manufacturing site. BDE 99 concentrations were identical or slightly higher than BDE 47 in most sediments (Allchin et al. 1999). Thesum 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 ( μ g/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 leach ing of PBDEs from landfills is minimal, movement of polymer particles containing pentaBDE, octaBD E, and decaBDE commercial mixtures within the landfill could lead to entry into leachate water of gr oundwater. 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 accoun ted 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). Th is is not a comprehensive list.

PBDEs were not identified in water samples coll ected 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 ag ricultural operations contaminated in the original episode in 1973–1974. The concen trations 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 Hex cel Corp and the White Chemical Company, the manufacturers of octabromo- and decabromobiphenyl , contained decabromobiphenyl and other lower brominated biphenyls down to hexabromobiphenyl (Di Ca rlo 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 FireMast er 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 hexabromobipheny l was used in commercial and consumer products in the United States, mostly in th e production of plastic products. Since the cessation of production of hexabromobiphenyl, all of these pr oducts, 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 materi als 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 Michig an 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 μ g/kg PBBs, whereas soil in r esurfaced cattle-exercise lots contained as high as 1,000–2,000 μ g/kg of PBBs (Fries 1985b).

PBBs have been identified in 5 soil and 3 sedim ent 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 m unicipal wastes with the disposal of consumer products. Solid waste from commercial production of oc taBDE is typically disposed in landfills (EPA 1995). The disposal of consumer products containing PBDEs is likely to increase worldwide due to rapidPBBs 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 m illion 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 μ g/kg dry weight; the levels of pentaBDE were high and cons istent, regardless of the region of origin. The concentration of decaBDE (BDE 209) varied wi dely among biosolids from different regions; the concentration of BDE 209 ranged from 84.8 to 4,890 μ g/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 si nce 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 fr om 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 de position and should not travel long distances in the environment. In water, PBBs ar e expected to absorb str ongly to suspended solids and sediment, and may bioconcentrate in aquatic organi sms. The volatilization of PBBs from water to air is not expected to be important due to attenuation by adsorption in the water co lumn. 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). Th e 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 dr y 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 part itioning of PBBs in water. Based on an estimated Henry's law constant of 3.9x10 -6 atm-m 3 /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 tr ansport 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 decabrom obiphenyl 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 tr ansport 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 bioconcentrati on 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 wa ter) 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 fa thead 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 hi gher degrees of bromination. A similar trend in BCFPBBs and PBDEs 397

8. POTENTIAL FOR HUMAN EXPOSURE

values for various PBBs was also obser ved in juvenile Atlantic salmon ( Salmo salar ). For example, the whole body BCF values determined for 2,6-dibr omobiphenyl, 2,4-dibromobiphenyl, 3,4-dibromobiphenyl, 2,5,4'-tribromobiphenyl, 2,2',4,5'-te trabromobiphenyl, 2,3',4',5-tetrabromobiphenyl, hexabromobiphenyl (unspecified c ongener), 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 ha ve very limited water solubility and are therefore not available to penetrate membranes (Zitko 1979).

PBBs are adsorbed strongly to soil, and the ad sorption increases with an in crease 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 adso rption. The presence of cer tain 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 groundwat er 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 PB Bs) are present at significant concentrations in soil as would happen at some contaminated sites (Griffi n 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 vi a 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 a nd dry deposition. Thus, in general, PBDEs are not expected to travel l ong 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 conge ners have been transported such long distancesPBBs and PBDEs 398

8. POTENTIAL FOR HUMAN EXPOSURE

from areas of emission. However, it is likely that th ey have been transported to remote regions as dust particles rather than in the vapor phase. In water, PBDEs are expected to ad sorb strongly to suspended solids and sediment, and bioconcentrate in aquatic or ganisms. The volatilization of PBDEs from water to air is not expected to be important due to attenuation by adsorption in the wate r column. In soil, PBDEs are adsorbed strongly and will be immobile. They ar e 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 en tirely 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. Watana be 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), whic h 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., pentaB DEs and hexaBDEs) may have the potential to be transported and would likely be fou nd 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 mixtur es of pentaBDE and decaBDE are 13.3 and <0.1 μ g/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 sed iments in the water column. Volatilization of PBDEs from water surfaces will be attenuated by adsorpti on, and thus is not an important fate process. Sediment-water partition coefficients (K p ) have been measured for several components of commercial pentaBDEs (Watanabe 1988). K p 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 K ow values have been measured for PBDEs as follows (Watanabe and Tatsuka wa 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

K ow values, log organic carbon-water partition coefficients (K oc ) 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 pr oducts 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, indicat ed little bioconcentration in carp (e.g., Cyprinus carpio )

with a bioconcentration factor of <4 after 8 week s of exposure (WHO 1994a). A bioconcentration study

was carried out with rainbow trout under static conditions. The concentration in the water was 20 μ g

14 C-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 ac cumulation 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 approximat ely 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 1x10 9 for BDE 99 and BDE 100, and ≈ 2.5x10 7 for BDE 28, ≈ 2.5x10 8 for BDE 47, and ≈ 1.6x10 8 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). Highe r brominated congeners (e.g., decaBDE) are rarely

detected in biota. This is a resu lt of their low solubility, high log K ow 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 w ould be expected due to their greater water solubility and relatively high K ow 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 conc entrations 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 tetr abromodiphenyl 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 collect ed 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 K ow 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 ve ry 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 wate r 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 transforma tion 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 numbe r 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 hydr oxyl (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 rela tionship 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 brominat ed PBBs in the atmosphere due to reaction with OH radicals are probably not important.

Hexa- and other higher brominated biphenyls are expect ed to be present in the particle-adsorbed state in the atmosphere. These PBBs photolyze in solution an d in soil (Hill et al. 1982; Ruzo and Zabik 1975; Trotter 1977). Since PBBs present in surface soil ar e 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 PBDE s 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 pent a-, 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 5x10 5 hydroxyl radicals per cm 3 and is based on a 24-hour day ofPBBs 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 underg o 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- a nd 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 (Ruz o 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, th e 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 main ly 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 FireM aster BP-6 (Robertson et al. 1983b).

The study of photolysis of PBBs in the aqueous phase is more relevant to natura l 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, phot olysis 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 hydroxylate d species (phenolic products) was found (Ruzo et al. 1976).

PBBs are not expected to undergo abiotic hydrolysis under environmental conditi ons 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 intermedia te by mixed bacterial

cultures obtained from PBB-contaminated river sedimen t (Kong and Sayler 1983). However, complete

mineralization was not observed for 2- a nd 3-bromobenzoate (K ong and Sayler 1984).

Although higher brominated biphenyls do not biode grade 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 absorb ed light up to 325 nm, which indicates that these compounds may be susceptible to photodegradation at e nvironmental 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 isPBBs and PBDEs 404

8. POTENTIAL FOR HUMAN EXPOSURE

successively debrominated by UV light to hexaBDE and th at 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 degr ade with a half-life of around 30 minutes while halflives for tetra-, penta-, hexa-, hepta-, and octabrom odiphenyl ethers were 12–16 days, 2.4 days, 1.2 days, 1.2 days, and 5 hours, respectively. The decompos ition products of decaBDE were identified to be PBDEs (with >6 bromine atoms per molecule) and po lybrominated 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 (H 2 O: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 minu tes) 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 fo und 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 ex amined 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 dissol ved in toluene was deposit ed on the solid substrate under a stream of nitrogen (to evaporate the solvent) a nd then desiccated to remove any residual toluene.

The adsorbed decaBDE on the solid s ubstrate was then inundated with the aqueous test solution, followed by irradiation for the duration of the test period. In all experiments, natural sunl ight (location, 40° 26’ N, 86° 55’ W) was used. The extent of degrada tion 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, sola r irradiation of decaBDE adsorbed onto humic acidsand indicated that the photolysis of decaBDE was sl ow. After 96 hours of exposu re 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 tub es 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 disa ppearance of decaBDE vs. the appearance of bromide ion) suggests that after the initial degradation of d ecaBDE, bromide ion was generated by the degradation of lower brominated diphenyl ether congener produc ts (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 phot odegradation of decaBDE adsorbed on glass tubes, which were filled with aqueous solutions (without humi c acid). The result of this test showed a much more rapid loss of decaBDE than found in the analog ous 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 presen t as either decaBDE or bromide ion, while the remaining 50% was possibly unidentified nona- and oc tabromodiphenyl 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 nonabromodi phenyl ethers under somewhat environmentally relevant conditions. Lower brominated di phenyl 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 degradati on 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 wa s allowed to evaporate, and then the sample was

reconstituted with water to resemb le natural conditions. BDE 209 was photolytically labile and formed debromination products in all matrixes studied. No na- to tetra- BDEs were formed as well as some PBDFs. The half-lives in toluene and on silica ge l 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 experime nts show that photolytic debromination of BDE 209 is a possible pathway for the formation of more bioa vailable, lower brominated PBDEs. However, the mostt commonly found BDEs in environmental samples (e.g., 2,2’,4,4’-tetrabromodiphenyl etherPBBs and PBDEs 406

8. POTENTIAL FOR HUMAN EXPOSURE

[BDE 47], 2,2’4,4’,5-pentabromodiphenyl ether [B DE 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 w as adsorbed on glass tubes filled with an aqueous solution and exposed to natural sunlight. Afte r 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 initiall y slow with the rate increasing after 24 hours while the disappearance of BDE 47 was initial rapid over th e first 24 hours. Using GC-MS, the authors concluded that 2,4,4’-tribromodiphenyl ether was bei ng 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. Th is study suggests that adsorbed pentabromodiphenyl ether congeners, like decaBDE, may undergo photoly sis 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 CO 2 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 CO 2 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 ha s 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 observe d. Thus, decaBDE was determined to be not readily biodegradable.

No data on biodegradation of pe ntaBDE and octaBDE commercial mi xtures under anaerobic conditions are available. An anaerobic degradation study was carried out with 2,2’,4,4’-tetrabromodiphenyl etherPBBs and PBDEs 407

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(BDE 47) using a mixture of 14 C labeled and unlabeled compound (EU 2003a). The test was carried out using a sediment-water (Schuykill River, PA) inocul um. 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 uni dentified products had been formed in samples taken after 32 weeks. From these results, it is clear th at BDE 47 has the potential to degrade slowly under anaerobic conditions (EU 2003a). Rayne et al. (200 3b) recently reported that 4,4’-dibromodiphenyl undergoes reductive debromination under anaerobi c 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 14 C-labeled decaBDE was studied over a period of 32 weeks (EU 2002). The test chambers consisted of 500 mL bo ttles 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

14 CO 2 and 14 CH 4 indicating that essentially no mineralization had occurred. GC-MS results showed no evidence for the formation of lower brominated conge ners 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 Ch ou 1981a, 1981b; Jacobs et al. 1976). Some degradation of an undefined congener of pentabromobiphenyl was observed when incubated in soil, but this degrada tion could not be definitely attribut ed to biodegradation (Jacobs et al. 1976). As discussed in Section 8.3.2.2, higher bromin ated PBBs may biodegrade in an anaerobic region of river sediment and possibly soil polluted with PC Bs and PBBs to form lower brominated products. Biodegradation of the photolysis products of hexa- and heptabro mobiphenyl 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 ofPBBs and PBDEs 408

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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 or iginal concentrations of PBBs were much lower, the rate of degradation was much lower. Principal degr adation products were 2,3',4,4',5-pentabromobiphenyl, 2,2',4,4',5-pentabromobiphenyl, and two unidentif ied 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 spr ead in field soil from Michigan (Jacobs et al. 1978). The authors provided no explanation for the differenc e in photoreactivity of PBBs in soils with and without manure. It is important to point out that, due to attenuation and scatteri ng of light, sunlight will not penetrate most soil beyond the surface layer. Ther efore, it can be concluded from these studies that although photolysis may be the onl y viable degradative process for PBBs in soil, photolysis will be limited to the surface layer of soil, and the rate of phot olysis will be very slow. PBBs incorporated into thermoplastics which were eventually buried at wast e 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 contamin ation. 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 upperm ost few millimeters of soil and will not impact sediment. Photolysis of PBDEs is possibily importa nt for land-applied sewage sludge contaminated with PBDEs. However, no information was available on th is possibility. Based on studies in water, most PBDEs are unlikely to biodegrade in soils or sed iment 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 (E U 2003a). However, information on the rate or extent of anaerobic biodegradation of BDE 47 has not b een found. No other information was found in the literature about the transformation and degradati on processes for PBDEs in soils and sediment.PBBs and PBDEs 409

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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 samp les 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 congene rs. 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 ca tastrophe of the early 1970's (see Section 8.1). Monitoring studies indicate that PBDEs are transported gl obally. 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 la te 1970s, with lower brom inated 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 dow nwind and crosswind from the White Chemical Company plant in Bayonne, New Jersey was 0.06 ng/m 3 (DeCarlo 1979).PBBs and PBDEs 410

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Polybrominated Diphenyl Ethers. In general, limited information is available in the literature reporting

the levels of PBDEs in ambient air. The concentr ations of PBDEs in air samples are summarized in Table 8-2. Atmospheric concentrations of PBDEs te nd 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 leve ls 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 measureme nts were averaged over a 3-year period between 1997 and 1999. The concentration of total PBDEs ranged from 5.5 pg/m 3 in rural environments to 52 pg/m 3 in urban air from Chicago, Illinois. The concentration of BDE 47 was 48 pg/m 3 observed near Chicago, Illinois. The average concentration of decaBDE at the remote and rural locations was <0.10 pg/m 3 for each of the years investigated. The averag e concentration of decaBDE in the particulate phase at the urban location ranged from 0.20 to 0.35 pg/m 3 (Strandberg et al. 2001).

Throughout the year of 1997, air samples were take n 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 ethe rs 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/m 3 at Hazelrigg and from 6 to 58 pg/m 3 at Stoke Ferry (de Wit 2002). PBDEs have also been measured in air samples taken from remote stations in the Arctic (e.g., Alert, Northw est 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/m 3 at Alert for the majority of the year; however, in July 1994, the concentration was 28 pg/m 3 . 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/m 3 , 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 w as found, although the limit of detection limit for decaBDE is much higher than for th e lower brominated diphenyl ethers.PBBs and PBDEs 411

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### Table 8-2. Concentrations (pg/m 3 

### ) of Several PBDEs in Air Samples

Location BDE 47 BDE 99 BDE 100 Σ PBDEs a 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 Σ PBDEs 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. 2001PBBs 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 fro m the Michigan Chemical Corporation plant in St. Louis, Michigan, to the Pine River during 1974–1977 ranged from <0.01 to 150 μ g/L (Hesse and Powers 1978). The concentrations of PBBs in effluents from White Chemical Company, Bayonne, New Jersey, and Hexcel Chemical Corporation, Sayerv ille, New Jersey, ranged from <0.2 to 210 μ g/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 μ g/L (Hesse and Powers 1978; Neufeld et al. 1977). 2,2',5,5'-Tetrabromobiphenyl and 3,3',5,5'-tetrabro mobiphenyl were qualitatively detected in water from Lake Ontario, and hexabromobiphenyl (unspecified congeners) was qualitatively detected in water from Lakes Ontario and Huron (Great Lake s 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 μ g/L, although the concentrations of PBBs in test wells outside the landfill ranged from <0.1 to 4.4 μ g/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 ~9 0% 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 th e 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 manu re on the fields ranged from 35 to 1,260 μ g/kg, while thePBBs and PBDEs 413

8. POTENTIAL FOR HUMAN EXPOSURE

concentrations in surface soil of control farms (t hat did not use contaminated manure) were <25 μ g/kg (Fries and Jacobs 1980).

Concentrations of PBBs in sediments upstream fro m the Michigan Chemical Corporation plant were below the detection limit (100 μ g/kg) with the exception of one sample (Hesse and Powers 1978). The concentration of PBBs in sediment from one upstream sample was 350 μ g/kg. Hesse and Powers (1978) explained that this higher value was due to contam ination by upstream currents during periods of waterlevel regulation at the St. Louis dam. The concentr ations 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 gradua l 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 sampl es from swamps and marshes adjacent to the White Chemical Company and Hexcel Chemical Corpor ation plants in New Jersey ranged from <10 μ g/kg to 4.6 mg/kg (DeCarlo 1979). A sludge sample from the di scharge treatment plant of the White Chemical Company contained 431 mg/kg of PBBs (DeCarlo 1979).

Polybrominated Diphenyl Ethers. N o 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. Tota l PBDE levels in these samples ranged from not

detected to 76 μ g/kg dry weight. 2,2’,4,4’,5-Pentabromodi phenyl ether (BDE 99) was the predominantcongener 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 th at concentrations of PBDEs in sediments are increasing. Burdens of PBDEs in sed iment 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 poin t source. DecaBDE (BDE 209) was the major congener detected at concentration ranging from 19 to 36 μ g/kg (ng/g) dry weight. Other congeners detected (inPBBs 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 Σ PBDEs 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, BD E 209, and other congeners (not specified)

dw = dry weight; ND = not detected

Sources: de Wit 2002; Hale et al. 2003PBBs and PBDEs 415

8. POTENTIAL FOR HUMAN EXPOSURE

decreasing order: 2,2’,4,4’,5-tetrabromodiphenyl ethe r [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-pentabromodiphe nyl ether [BDE 100]) were less than 5 μ g/kg (ng/g) dry weight. PBDEs were above the detection limit (i.e., 0.5 μ g/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 BD E 100. The maximum concentration detected in

sediment was 52.3 μ g/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 μ g/kg (ng/g) dry weight of pentaBDE.

In Japan, tetra-, penta-, hexa-, and decabromodipheny l 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 ( μ g/kg) dry weight. The concentra tion of decaBDE (BDE 209) ranged from <25 to 11,600 ng/g ( μ g/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 ( μ g/kg) dry weight for BDE 47, not detected to 5.5 ng/g ( μ g/kg) dry weight for BDE 99, and not detected to 510 ng/g ( μ g/kg) dry weight for BDE 209. The concentration of PBDEs in suspended particulat e matter ranged from not detected to 9 ng/g ( μ g/kg) dry weight for BDE 47, not detected to 23 ng/g ( μ g/kg) dry weight for BDE 99, and not detected to 4,600 ng/g ( μ g/kg) dry weight for BDE 209 (de Boer et al. 2000). The concentration of several

brominated flame retardants was measured in sedimen ts collected from the mouths of major European rivers (de Wit 2002). Elevated levels of BDE 47 a nd BDE 99 were found in Humber and Mersey rivers (Great Britain). In two rivers of the Netherla nds, the sum of BDE 47 and BDE 99 ranged 1.61 to 13.1 ng/g ( μ g/kg) dry weight. The highest hexaBDE levels (as BDE 153) were found in the river Seine (France), three rivers in the Netherlands, and the rive rs Schelde (Belgium), Forth (Great Britain) and Ems (Germany); the concentration of BD E 153 ranged from 0.013 to 0.056 ng/g ( μ g/kg) dry weight in these

sediments. The concentrations of decaBDE were hi ghest in sediment from the Seine, ranging from 2.4 to 3.9 ng/g ( μ g/kg) dry weight. The concentrations of decaB DE in River Mersey (Great Britain), Schelde and River Liffey (Ireland) ranged from 34 to 1,800 ng/g ( μ g/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 ( μ g/kg) dry weight (Nylund et al. 1992).

A well-studied sediment core collected from the southe rn part of the Baltic Sea proper was analyzed for PBDEs and a number of organochlorin e contaminants (Nylund et al. 1992). The retrospective temporalPBBs 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 fresh water lake in Germany, the Wa dden 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 15 4, 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 show s increasing levels of BDE 47 starting in the 1940s (range, 0.02–0.18 ng/g dry weight) and increasing leve ls 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-o ff 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 exis tence 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 unknow n sources of PBDEs produced in nature.

8.4.4 Other Environmental Media

Polybrominated Biphenyls.

Food. Although the agriculture episode in Michigan invol ving 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 fr om 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 pr oducts, the levels of PBBs in consumer products showed a steadyPBBs and PBDEs 417

8. POTENTIAL FOR HUMAN EXPOSURE

decline. For example, in 1975, among 18 milk sa mples, 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. 197 8). Although 95% of 1,430 meat samples collected in Michigan in 1976 contained detectable PBBs, onl y 1 sample contained >0.6 mg/kg, and a market basket survey in Michigan showed detectable PBBs in only 1/102 m eat 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 upstr eam (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 f illets). There was no apparent change in PBB

concentrations in fish between 1974 and 1976 (Hesse and Powers 1978). P BBs 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 tr ibutaries of Lake Huron (including Pine River) and Lake Superior were determined (G reat Lakes Water Quality Board 1989; Jaffe et al. 1985). PBBs were detected in the concentration range of 15–15,000 μ g/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 tr out 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 fro m 189 to 2,083 pg/g wet weight and from 42 to 633 pg/g wet weight, respectively. Several other congeners we re also detected in these lake trout samples (see Table 8-4).

In German rivers, elevated levels of nona- and oc taBBs 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 ( μ g/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 ( μ g/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 ( μ g/g) and not detected to 1.8 mg/kg ( μ g/g), respectively (Hesse and Powers 1978). Three bottlenose dolphinsPBBs and PBDEs 418

8. POTENTIAL FOR HUMAN EXPOSURE



### Table 8-4. Mean Concentr ations of Nine PBB Congen ers 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. 2002PBBs 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 μ g/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 ( μ g/g), respectively (Kaiser et al. 1980). Twenty-two othe r specimens did not contain detectable levels (<0.03 mg/kg [ μ g/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 contai ned PBBs in the concentration range of 0.02– 0.25 mg/kg ( μ g/g) wet weight (Heinz et al. 1983, 1985).

White-tailed sea eagles collected from the Baltic Sea cont ained 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 Swed ish 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 μ g/kg (ng/g) wet weight from harbor seal s collected from the North Sea (decaBB <1 μ g/kg [ng/g] wet weight). In whitebeaked dolphins from the Nort h Sea, the concentration of hexa-, penta-, and decaBBs were 13, 8.3, and <0.9 μ g/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 μ g/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 repos itory of PBBs in the body, and concentrations in fat can provide an index of body burde ns 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 ca n 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 μ g/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 μ g/kg (ng/g) (Lewis and Sovocool 1982). PBB levels in the adipose tissu es 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 ( μ g/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 fro m the Great Lakes, were measured (Anderson et al. 1998). When the data were stratified by lake, on averag e, 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 epis ode in 1973–1974, the median PBB concentrations in blood of exposed adults and children in farms we re 0.014 and 0.035 mg/kg (14 and 35 ppb [ng/g]), respectively, compared to corresponding median con centrations 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 no nquarantined 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 ove r 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 le vels of PBBs in human tissues. Levels of PBBs were highest in the part of state in which the episode occurred (median: adipose tissue, 500 μ g/kg (ng/g); serum, 1.7 μ g/L) and were lowest in the upper peninsula (median: adipose tissue, 15 μ g/kg (ng/g); serum,

0.2 μ g/L), farthest from the source of contamination. Le vels in the rest of the state were in between (median: adipose tissue, 240 μ g/kg [ng/g]; serum, 0.9 μ g/L) (Wolff et al. 1982). The estimated concentration ratio of PBBs in adipose tissue ove r 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 wome n (Brilliant et al. 1978). The kinetics of fatPBBs and PBDEs 421

8. POTENTIAL FOR HUMAN EXPOSURE

metabolism in lactating women seems to alter PBB pa rtitioning. 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 female s were 0.10–0.14 to 1 an d 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. Howeve r, the concentration of PBBs in feces represented a minor proportion of the total body burden, indicating a sl ow 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 fa rm quarantined by the Michigan Department of Agriculture, were recruited from July 1976 to D ecember 1977, and were followed up annually from 1978 through 1993. Median serum PBB concentra tions were 2 ppb (n=290; range=0.5–419 ppb [ μ g/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 al so has been studied (Murata et al. 1977; Zabik etal. 1978). Spray-drying reduced PBB levels in whole and skim milk, whereas pasteurization, freezedrying, aging of cheese, and condensation were not e ffective 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 P BBs in Michigan Residents

Tissue Subject(s) Mean/median concentration a Year Reference Serum Exposed farm workers 14 μ g/L 1976 Stross et al. 1979 Chemical workers 48 μ g/L No data Stross et al. 1981 Chemical workers 1.1–

 1,000 μ g/L 1976 Anderson et al. 1978d Exposed farm workers BDL to 1,000 μ g/L 1976 Anderson et al. 1978d

Residents from quarantined farms 26.9 μ g/L 1976–1977 Landrigan et al. 1979 Residents from non-quarantined farms 3.5 μ g/L 1976–1977 Landrigan et al. 1979

Farm product consumers 17.1 μ g/L 1976–1977 Landrigan et al. 1979 Chemical workers and families 43.0 μ g/L 1976–1977 Landrigan et al. 1979 Control group 3.5 μ g/L 1976–1977 Landrigan et al. 1979 General population (lower peninsula) 1.9 μ g/L 1978 Wolff et al. 1982 General population (upper peninsula) 0.2 μ g/L 1978 Wolff et al. 1982

General population (remainder of state) 0.9 μ g/L 1978 Wolff et al. 1982

Chemical workers 25.4 μ g/L No data Eyster et al. 1983 Farm and other workers 5.4 μ g/L No data Eyster et al. 1983 Mothers from lower peninsula 26.2 μ g/L 1976–1977 Landrigan et al. 1979 Exposed mothers from farms 3.4 μ g/L No data Eyster et al. 1983 Non-pregnant women from exposed farms 3.1 μ g/L No data Eyster et al. 1983

Exposed women enrolled in the Michigan Department of Public Health registry 2 μ g/L 1993 Henderson et al. 1995

Cord serum Exposed mothers from lower peninsula 3.2 μ g/L 1976–1977 Landrigan et al. 1979

Mothers from lower peninsula <1.0 μ g/L No data Eyster et al. 1983 Blood plasma Workers from quarantined farms 14 μ g/L 1974 Humphrey and Hayner 1975 Children from quarantined farms 35 μ g/L 1974 Humphrey and Hayner 1975 Adults from non-quarantined farms 3 μ g/L 1974 Humphrey and Hayner 1975 Children from non-quarantined farms 6 μ g/L 1974 Humphrey and Hayner 1975 Placenta Exposed mothers <1 μ g/L No data Eyster et al. 1983 Breast milk Exposed mothers 370 μ g/kg (fat basis) No data Eyster et al. 1983

Exposed mothers from lower peninsula 3,614 μ g/kg (fat basis) 1976–1977 Landrigan et al. 1979

Mothers from lower peninsula 68 μ g/kg (fat basis) 1976 Brilliant et al. 1978

Mothers from upper peninsula <44 μ g/kg (fat basis) 1976 Brilliant et al. 1978PBBs and PBDEs 423

8. POTENTIAL FOR HUMAN EXPOSURE



### Table 8-5. Tissue Levels of P BBs in Michigan Residents

Tissue Subject(s) Mean/median concentration a Year Reference Adipose tissue Population of lower peninsula 500 μ g/kg 1978 Wolff et al. 1982

Population of upper peninsula 15 μ g/kg 1978 Wolff et al. 1982 Population of rest of the state 240 μ g/kg 1978 Wolff et al. 1982 Chemical workers 9,330 μ g/kg No data Brown et al. 1981 Farm residents from lower peninsula 3,940 μ g/kg No data Brown et al. 1981 Farm residents from lower peninsula 3,260 μ g/kg 1976 Stross et al. 1979 Chemical workers 12,820 μ g/kg No data Stross et al. 1981 Workers from quarantined dairy farms 12,500 μ g/kg 1974 Humphrey and Hayner 1975 Pregnant females from lower peninsula 400 μ g/kg No data Eyster et al. 1983

Chemical workers 5,290 μ g/kg No data Eyster et al. 1983 Farm and other workers from lower peninsula 1,650 μ g/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 limitPBBs 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 ch ickens raised in two different regi ons 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 Dakot a was 1.7 ng/g whole weight. In the United States, chickens fed ball clay and chickens bought in the gr ocery 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 weig ht 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. Th e concentrations of PBDEs in spinach, potato, and carrot were 134, 47.6, and 38.4 pg/g fresh weight, r espectively. 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 anal yzed. 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 co ntamination 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 so ld 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 hum ans. The concentrations of PBDEs in seafoods from the Inland Sea of Japan we re 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-pe ntabromodiphenyl ether (BDE 100); BDE 153; and 2,2’,4,4’,5,6’-hexabromodiphenyl ether (BDE 154) we re 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 concen trations of several PBDE concengers in omnivorous and vegetarian diet samples from the United Kingdom. Median concentrations of BDE 47PBBs 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 omnivorou s diet samples were 66.8, 63.8, 10, 20, and 20 pg/g dry weight, respectively. In vegetarian samples, me dian concentrations of BDE 47 (2,2’,4,4’-tetraBDE), 99 (2,2’,4,4’,5-pentaBDE), 100 (2,2’,4,4’,6-p entaBDE), 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 f ound 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 co nsistent, regardless of the region of origin. The concentration of decaBDE (BDE 209) va ried 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 (Vir ginia) were collected and analyzed for PBDEs (Hale et al. 2002). Congener patterns suggestive of bot h 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 s ource of pentaBDE discharged to this plant, the distribution pattern for lower brominated congene rs 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 concen trations 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 thermoplastic s (e.g., computers). Concentrations of PBDEPBBs and PBDEs 426

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### 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 Σ PBDEs a 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 Σ PBDEs 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. 2002PBBs and PBDEs 427

8. POTENTIAL FOR HUMAN EXPOSURE

congeners were 107–174 μ g/kg dry weight basis for 2,2’,4,4’-tetrabromodiphenyl ether (BDE 47), 51.1–

74.1 μ g/kg dry weight basis for 2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100), 155–293 μ g/kg dry weight basis for 2,2’,4,4’,5-pentabromodiphenyl ether (BDE 99), 42.0–53.5 μ g/kg dry weight basis for 2,2’,4,4’,5,5’-hexabromodiphenyl ether (BDE 153), 219–305 μ g/kg dry wt basis for 2,2’,4,4’,5,6’-hexabromodiphenyl ether (BDE 154)/PBB-153, and 1,330–2,660 μ g/kg dry wt basis for decaBDE (BDE 209).

Levels of PBDEs were found to be si milar 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 leve ls of PBDEs are increasing in freshwater organisms with higher concentrations near point sour ces. 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-pentab romodiphenyl 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). BD E 47, BDE 99, BDE 153, and BDE 154 were primary congeners. From the Lake of O zarks, 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). Th e 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- a nd 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. Te traBDE 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- toPBBs 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 Σ PBDEs a 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. 1999PBBs and PBDEs 429

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### 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 Σ PBDEs a 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 Σ PBDEs 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 weightPBBs and PBDEs 430

8. POTENTIAL FOR HUMAN EXPOSURE

heptaBDEs with combined concentra tions 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, ma y 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 Onta rio show a dramatic increase in total PBDE concentrations over time (Luross et al. 2000). At 50 fresh water sites in Vi rginia, muscle samples from 253 fish samples were collected and analyzed for PB DEs (Hale et al. 2000, 2001b). Approximately 85% of the samples contained BDE 47, the predomin ant 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 id entified 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 [B DE 181] and 2,2’,3,4,4’,5’,6-heptaBDE [BDE 183]), lesser amounts of 2,3,3’,4,4’,5,6-hept aBDE (BDE 190), and two hexaBDEs (BDE 154 and BDE 153). Possible sources for the heptaBDE congene rs were not obvious since none of the commercial mixtures are known to contain these congeners. Three po ssible 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 sa mples 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-pentabro modiphenyl 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 Vä ttern) 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 Arc tic 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 Da lslands 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 th ese samples was similar to the technical mixture, Bromkal 70-5DE, with equa l quantities of both BDE 47 and BDE 99. The levelsPBBs and PBDEs 431

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### 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 Σ PBDEs a 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 Σ PBDEs 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 weightPBBs 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 locat ed (Andersson and Blomkvist 1981). These textile industries have used PBDEs in the production of te xtiles. 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 we re 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 to tal 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 BD E 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). Leve ls of BDE 47 ranged from <20 to 1,700 ng/g lipid in eels ( Anguilla anguilla ) from Dutch rivers and lakes (at 10 locatio ns); 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 susp ected to contain PBDEs. Levels of BDE 47, BDE 99, pentaBDE (as technical mixture DE-71), a nd 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

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Saltwater Fish. No identifiable temporal trends were found for PDBE levels in marine aquatic species. Spatial trends show higher levels of lower brominated BDE congene rs 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 ar e 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 congene rs (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. Deca BDE 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 BD E congener levels were 1,188 and 4,147 ng/g wet weight for the two sample s. Likewise, wild salmon from four locations in Canada were analyzed for BDE congeners. Levels we re 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 con centrations in skipjack tuna from Asian offshore waters, off-Seychelles, off-Brazil, and open seas we re determined for samples collected during 1996– 2001 (Ueno et al. 2003). The concentration of to tal BDEs in muscles tissues ranged from not detected (<0.05 ng/g lipid) to 53 ng/g lipid. The concentrati on of the highest congener in muscle tissues, BDE 47, ranged from <0.1 to 15 ng/g lipid. BDE 99, BDE 1 00, BDE 153, and BDE 154 also were detected; BDE 209 was below the detection limit (<5.0 ng/g lipid ) for these samples. Sa mples 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 south east Asia had the highest. Fall-caught herring ( Clupea harengus ) muscle from five sites along the Swedish co ast was analyzed for BDE 47, BDE 99, and BDE 100; the combined concentration of these thr ee 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 diffe rent 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 ha d 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 w as not evident for herring. BDE 47 was thePBBs and PBDEs 434

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

Sample type Location BDE 47 BDE 99 BDE 100 Σ PBDEs a 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 dat a 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. 2000PBBs and PBDEs 435

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

Sample type Location BDE 47 BDE 99 BDE 100 Σ PBDEs a 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 Σ PBDEs 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. 2003PBBs 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 herri ng 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, resp ectively, in whole-body composites (Burreau et al. 1999). In another study, the levels of BDE 47, BD E 99, and BDE 100 were determined in muscle, ripe eggs, and blood plasma from Baltic sa lmon (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; bl ood, 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 dr y 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 le vels 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 envir onmental 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 ma rine fish and shell fish in Japan. TetraBDE and pentaBDE levels ranged from 0.1 and 17 ng/g fresh we ight, with tetraBDE being the major congener. decaBDE was also detected in a mussel sample from Osaka Bay (at 1.4 μ g/kg wet weight). Recently, Japanese market fish were analyzed for PBDEs. Th e 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 a nother 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/gPBBs and PBDEs 437

8. POTENTIAL FOR HUMAN EXPOSURE

fresh weight, with BDE 47 as the predominant congene r. 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 yellowt ails 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 dr y weight and from 0.3 to 1.6 ng/g dry weight, respectively. BDE 153 and BDE 209 were not detect ed. 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 prim ary congener detected was BDE 47. The combined

concentration of BDE 47 and BDE 99 was a pproximately 100–350 ng/g lipid weight.

Marine Animals. In marine animals, temporal trends show increasing levels of lower brominated BDE congeners with higher levels f ound near human-populated areas. In all marine animal studies, the congeners profile show the highest levels for BDE 47. The concentra tions 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 p ast decade, with current levels among the highest reported for this species (She et al. 2002). The concentration of to tal 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 blu bber 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 hi gher 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 BD E 100) (Jansson et al. 1993); male grey seals had 280 ng PBDE/g lipid weight (Andersson a nd Wartanian 1992). Male ringed seals ( Pusa hispida ) from the Baltic Sea had 320 ng PBDE/g lipid weight (Anderss on and Wartanian 1992). Baltic gray and ringed sealPBBs and PBDEs 438

8. POTENTIAL FOR HUMAN EXPOSURE



### Table 8-10. Concentrations (ng/g Li pid Weight) of Several PBDEs in Marine Animals

Sample type Location BDE 47 BDE 99 BDE 100 Σ PBDEs a 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 (f w)* 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 1992PBBs and PBDEs 439

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### Table 8-10. Concentrations (ng/g Li pid Weight) of Several PBDEs in Marine Animals

Sample type Location BDE 47 BDE 99 BDE 100 Σ PBDEs a 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 dat a 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 Σ PBDEs 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 weightPBBs 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 lipi d 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; Hagl und 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 (Ande rsson and Wartanian 1992). She et al. (2000) analyzed the concentration of BDE 47, BDE 99, a nd BDE 153 in harbor seal from the San Francisco Bay area (She et al. 2000). Mean concentrations fo r 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, Nort hwest 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 tota l PBDE levels (sum of tri- to hepta-congeners) were 350–2,300 ng/g lipid weight; 2,2’,4,4’-tetrabro modiphenyl ether (BDE 47) was found at the highest levels in these samples (range, 50–1,200 ng/g lipid we ight) (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 (B DE 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 dete rmined 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 pilo t 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 PB DEs and BDE 47 were 15 and 10 ng/g lipid weight, respectively (Stern and Ikonomou 2000). Between 1982 a nd 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 th e 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; a nd 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 conge ners 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’-heptaBD E (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 Michig an colonies and in Toronto Harbor, Lake OntarioPBBs 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 c oncentrations of PBDEs in common guillemots ( Uria aalge ) collected in 1979–1981 from the Baltic and Nort h 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 a nd 190 ng/g for BDE 99 in 1990 (Sellström 1999; Sellström et al. 1993). Between 1992 and 1997, PBDE le vels started to decrease statistically. In 1997, the PBDE level (sum of BDE 47, BDE 99, and BD E 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 determ ining the human body burden of these chemicals. Increasing levels of lower brominated PBDEs have b een measured in blood and breast milk in temporal

trend studies. Individuals who cons umed 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 concentr ations 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 flui ds are a factor of 10–100-fold higher for individuals living in the United States compared to individuals li ving 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 (P atterson et al. 2000; Sjödin et al. 2001). Concentrations of these congeners were similar to t hose 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 sa mples (Patterson et al. 2000). Serum samples werePBBs and PBDEs 443

8. POTENTIAL FOR HUMAN EXPOSURE



### Table 8-11. Concentrations (ng/g Lipi d Weight) of Several PBDEs in Human Blood Samples

Sample type Location BDE 47 BDE 99 BDE 100 Σ PBDEs a 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 Σ PBDEs 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. 2001PBBs and PBDEs 444

8. POTENTIAL FOR HUMAN EXPOSURE



### Table 8-12. Concentrations (ng/g Lipi d Weight) of Several PBDEs in Human Adipose Tissue Samples

Sample type Location BDE 47 BDE 99 BDE 100 Σ PBDEs a 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 dat a 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 dat a No data Meneses et al. 1999

Human adipose tissue Japan 459 118 250 1,288 No data Choi et al. 2003

a Σ PBDEs 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 detectedPBBs and PBDEs 445

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### Table 8-13. Concentrations (ng/g Lipi d Weight) of Several PBDEs in Human Breast Milk Samples

Sample type Location BDE 47 BDE 99 BDE 100 Σ PBDEs a 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 Σ PBDEs 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 (a ged 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 fro m 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. Thei r 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 le vels 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-re lated 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 gene ral, the PBDE congener profile found in human serum was similar to that detected in environmental sampl es, 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 bl ood-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 PB DEs 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 me asurable quantities of BDE 28, BD E 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) increasedPBBs 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 concentra tion 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 (B ergman et al. 1999; Sjödin et al. 2000). High fish intake groups of Swedish and Latv ian 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-d isassembly plant, workers in a computerized office, and a control group) (Sjödin et al. 1999b). The medi an 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 com puter 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 concentrat ions of BDE 153, BDE 18 3, and BDE 209 in these samples were found to be approximately 5 times highe r than the control and office workers in the Sjödin et al. (1999b) study. The median concentration fo r 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. Conn ections were observed between fish consumption and serum concentrations for conge ners BDE 47, BDE 153, and BDE 183, and between worktime at the computer and c ongeners BDE 153 and BDE 183.

DecaBDE, as well as hexa- through nonaBDE, h as 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 a nd 0.7 ng/g fat. In the late 1990s, breast adipose samples collected in northern California contained qua ntifiable 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). MeanPBBs 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 leve ls (86 ng/g lipid) were the highest human levels reported to date. Recently, Petreas et al. (2003) expand ed 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 concen trations of BDE 47 in ad ipose tissues samples taken from California women were 28.9 ng/g lipid. In th e 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 we re 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 concentr ations ranging from 2 to 2.4 ng/g lipid weight, BDE 99 concentrations of 1.6 ng/g lipid weight, BD E 100 concentrations of 0.1 ng/g lipid weight, and BDE 153 concentrations ranging from 1 to 1.3 ng/g lipi d 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 occu rred 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 c oncentrations 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 Hexa BDE were 0.93 and 1.83 ng/g lipi d weight, respectively (Meneses et al. 1999).

Human Milk. Recently, Schecter et al. (2003b) reported th e 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 clinicPBBs and PBDEs 449

8. POTENTIAL FOR HUMAN EXPOSURE

in Dallas, Texas, both in the year 2001. The medi an concentration of the sum of PBDE congeners was 34.0 ng/g lipid. The predominant congener was BDE 47 (1 8.4 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 concentra tions 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 wa s detected in 7 out of 47 samples with a maximum concentration of 8.24 ng/g lipid. PBDE levels in br east milk from this study were similar to levels found in U.S. blood and adipose tissue lipid from Califor nia 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 tem poral trends of PBDE concentrations in pooled breast milk samples from mothers in Stockholm, Swed en. Between 1972 and 1997 , the concentration of PBDEs in human breast milk increased, with a doubli ng 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 congene rs) 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-pent aBDE (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 of2,4,4-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 sam ples (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 samp les had total PBDE concentrations (sum of BDE 28, BDE 47, BDE 99, BDE 100, BDE 153, and BDE 154) ra nging from 0.66 to 2.8 ng/g lipid weight (Ohta etPBBs and PBDEs 450

8. POTENTIAL FOR HUMAN EXPOSURE

al. 2002a). Women who consumed fish had a somewh at higher concentration of total PBDEs (range, 1.4– 2.8 ng/g lipid weight) compared to women who at e 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 en vironment has not yet been explained. Anthropogenic sources of these compounds have not been found. Te tra- 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 bi obromination have been suggested as the origin for these compounds. Naturally produced methoxy-tetrabro minated 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 fi ve 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 observe d, 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 histori cal releases. For people residing in the lower peninsula of Michigan, especially in the immediate vici nity of the PBB contaminated areas of this region, exposure to PBBs may still be occurring today. Howe ver, environmental levels have decreased since the 1970s and current exposure, if any, w ill be at low levels. For other regi ons 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 cons umer 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 directPBBs 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 da ta indicate that low level exposures to PBBs were limited to the population within the state of Michig an (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 plas tics (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 bur den 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, worker s 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 insuffi cient to reach definitive conclusions. Humans appear to be exposed to lowe r 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 PBDE s such as textiles or polymers. Inhalation exposure could occur from outgassing of PBDEs from electrical a ppliances 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 le vels 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-pentaBDEPBBs 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 ge neral, levels of decaBDE in human tissues and body fluids are negligible which indicates that direct exposure to d ecaBDE 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/m 3 in the United States (Dodder et al. 2000a). Typically, lower brominat ed congeners (e.g., 2,2’,4,4’-tetraBDE [BDE 47]) are predominant which indicates low levels of exposur e to these congeners by the general population. Levels of higher brominated congeners (e.g., decaB DE) tend to be below the limit of quantitation. In Sweden, indoor air concentrations of PBDEs in l ecture 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 PBDE s 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 poly mers is dependant on the migration ability of the PBDE molecule through the polymer matrix to th e 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 correla tion between PBDE concentrations in indoor air and both the number of electrical appliances and th e number of chairs containing polyurethane foam. Concentrations of tetra- and pent abrominated 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/m 3 ), 99 (852 pg/m 3 ), and 100 (217 pg/m 3 ), respectively. In this study, the median lowe r bound estimates of inhalation exposure (i.e., where a congener is below the detection limit, the con centration 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 ma jor route of exposure in humans (Lindström 1999). Consumption of fish has been associated with elev ated 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 μ g of pentaBDE and 0.3 μ g of total PBDEs from fish that is ingested by humans daily (WHO 1994a). The fish of greatest concern to human s are bottom feeders like ca rp 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 medi an 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 hi gher 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). Occupa tional 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/m 3 (1.25 ng/m 3 ) for 2,2’,4,4’-tetraBDE (BDE 47), 4.6 pmol/m 3 (2.6 ng/m 3 ) for 2,2’,4,4’,5-pentaBDE (BDE 99), 6.1 pmol/m 3 (3.93 ng/m 3 ) for 2,2’,4,4’,5,5’-hexaBDE (BDE 153), 26 pmol/m 3 (18.8 ng/m 3 ) for 2,2’,3,4,4’,5’,6-heptaBDE (BDE 183), and 38 pmol/m 3 (36.5 ng/m 3 ) 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/m 3 ). Concentrations of PBDEs in the office (range, <0.002–PBBs and PBDEs 454

8. POTENTIAL FOR HUMAN EXPOSURE

0.09 ng/m 3 ) 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 medi an 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 com puter 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 concentrat ions of BDE 153, BDE 18 3, and BDE 209 in these samples were found to be approximately 5 times highe r than the control and office workers in the Sjödin et al. (1999b) study. The median concentration fo r 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 n onaBDE congener were also detected.

8.6 EXPOSURES OF CHILDREN

This section focuses on exposures from conception to matu rity 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 ma y 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. Child ren 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 u se the judgment of adults to avoid hazards (NRC 1993).PBBs and PBDEs 455

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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 formul a milk (see Section 8.4.4). Exposure of neonates is evident due to the presence of lower brominated BD Es (e.g., 2,2’,4,4’-tetraBDE [BDE 47]) in breast milk and placental tissue. The PBDEs detected in breastmilk are the tri- to hexaBDEs, but not the heptato decaBDEs (LaKind and Berlin 2000), which are th e 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 Unite d States. The median concentration of the sum of PBDE congeners was 34.0 ng/g lipid with BDE 47 (1 8.4 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 we re 10–100 times greater than human tissue levels in Europe (Schecter et al. 2003b). Levels of lower brom inated 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 y ear (Lindström 1999). The concentrations of lower brominated BDEs in breast milk are increasing exponentia lly from about 300 pg/g lipid in 1976 to about 4,000 pg/g lipid in 1997 in Swedish women (Nor ön 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-pent aBDE [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 Un ited 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 an d 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 le vels 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 197 5–2002. 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,PBBs and PBDEs 456

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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 plast ics are probably no longer in circulation. At the present time and in the near future, populations potenti ally 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 li fetime 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 contam inated meat and dairy pr oducts obtained from herds grazing over contaminated soil and consumption of fish from nearby contaminated streams. PBB contamination has triggered the issuance of one human h ealth advisory in the state of Michigan. As of September 30, 1993, recreational and subsistence fi shermen 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 le vels of PBBs associated with dietary intake (EPA 1993). The body

burden for PBBs in residents of contaminated ar eas 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 1 994a). Other populations with high exposure levels to PBDEs involve occupational exposures (see Secti on 8.6). No other information was located that identified specific populations with higher exposure levels to PBDEs.PBBs and PBDEs 457

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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 P BBs and PBDEs are available. Where adequate information is not available, ATSD R, in conjunction with NTP, is re quired 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 substan ce-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 f illed. In the future, the id entified data needs will be evaluated and prioritized, and a substance-sp ecific 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 physi cal and chemical properties of hexabromobiphenyl are

available relative to octabromo- a nd 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 hexabr omobiphenyls. More importa ntly, very limited data are available on the physical and chemical properties for the individual congene rs of hexabromo-, octabromo-, and decabromobiphenyl. The absence of such important data as K oc , vapor pressure, and Henry’s law constant, is a major impediment in the predicti on 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 K oc , vapor pressure, and Henry’s law constant, are necessary fo r the prediction of the environmental fate and transport of PBDEs.PBBs and PBDEs 458

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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 u se of PBBs are available (Neufeld et al. 1977). In the past, PBB-containing plastic was used in c onsumer 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 me dia 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 fro m production plants and disposal of PBB-containing plastics in landfills (Neufeld et al. 1977). Althoug h 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 th e efficiency of incineration as a method of disposal of PBBs present in the neat form in industrial wast es or in plastics. Environmental regulations regarding the manufacture and disposal of PBBs have been est ablished (EPA 1988a). Ac cording 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 in formation to the EPA. The Toxics Release Inventory (TRI), which contains this information for 1999, becam e 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 retardan ts in thermoplastics at levels ranging from 5 to 30% by weight (EU 2001). The commercial pentaB DE 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 almo st exclusively as a flame retardant for acrylonitrilebutadiene-styrene (ABS) terpolymers used in co mputer casings and monitors (ENVIRON 2003b). The commercial decaBDE product is an additive flame retard ant 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-containi ng consumer products is described as transfers to disposal (landfill), recyclin g, energy recovery (inciner ation), or publicly owned treatment works (POTWs) (Darnerud et al. 2001). More information for PBDEs is needed on import/export, release, and disposal. In particul ar, the mechanism by which PBDEs are leaving the products in which they are used and entering the environment is not understood. Possibilities includePBBs and PBDEs 459

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disinigradtion of plastic products in particulates contaminated with PBDEs or volatilization of PBDEs from the plastic itself (Hites 2004). In the United St ates, 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 fr om incineration of PBDE-containing plastics would be useful in determining expos ure 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 (Ruz o and Zabik 1975; Ruzo et al. 1976), but there is less certainty about the importance of photolysis of PBBs in wate r (Norris et al. 1973; Ruzo et al. 1976). PBBs will partition from the aquatic phase to sediment and susp ended 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; O pperhuizen 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 pa thways (Jacobs et al. 1978). The translocation of PBBs from soil to upper parts in plants was not obser ved, 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 P BBs 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 absorbedPBBs and PBDEs 460

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light up to 325 nm, which indicates that these compounds may be susceptible to photodegradation at environmental wavelengths. However, the importanc e 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 studi es have been done on biodegradation of PBDEs inthe environment under both aerobic and anaerobic conditions, especially studies investigating dehalogenation mechanisms (EU 2002, 2003). More st udies 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-specifi c 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 c ontaminated 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 bioavailab ility of higher brominated biphenyls is expected to be even lower than the less brominated biphenyls du e 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 pers istent 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 g ood data exist on the adsorption of PBDEs commercial mixtures and congeners with the exception of d ecaBDE (BFRIP 2002). Studies on many persistent halogenated aromatic compounds clear ly show that they become progessively less bioavailable with time (Alexander 2000). Often, three-fourths or more of the concentration of such compounds is notPBBs and PBDEs 461

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bioavailable. Information on the possibility of the ve ry 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 da ta. Brominated biphenyls with

bromine substitution 6 or less will bioconcentrate in a quatic organisms (Gobas et al. 1989; Norris et al. 1973; Opperhuizen et al. 1985; Zitko 1979; Zitko and Hu tzinger 1976). PBBs are preferentially stored in the adipose tissue of animals (Kimbrough 1987). Althoug h PBBs have been detected in fish-eating birds and predatory animals from the consumption of cont aminated 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, wh ich results in bioconcentration (see Section 8.4.4; Hardy 2002b). Congener components of the pentaB DE commercial product tend to bioconcentrate to

different extents. DecaBDE and octaBDE comme rcial products appear to not bioconcentrate, bioaccumulate, or biomagnify (Hardy 2002b). The lim ited existing data indicate that lower brominated BDE congeners (e.g., pentaBDE commercial mixtures) bioaccumulate in aquatic and terrestrial food chains and biomagnify in predat ors 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 bioaccu mulation 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 Powe rs 1978; Shah 1978). No data on the level of PBBsPBBs and PBDEs 462

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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 obta ined on levels of PBBs in the environment can be used in combination with the known body burden of PBBs to assess the poten tial risk of adverse health effects in populations living in the vicinity of hazardous waste sites.

Polybrominated Diphenyl Ethers. Information on the relative importa nce 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 ar eas, as well as areas near hazardous waste sites and incinerators are needed. Due to the hydrophobic natu re of PBDEs, this class of compounds has not been detected in water to any significant extent. BDE 47 w as detected at low concentrations in Lake Ontario surface waters (Luckey et al. 2001). Although levels ar e predicted to be low, monitoring data on PBDE concentrations in finished drinking water nationwid e 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 hi gher 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 concentr ations of PBDEs in foods, collected using a marketbasket approach, are needed to de termine concentrations of PBDEs in foods consumed by the general population. Data on the PBDE concentrations in food s 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. Monito ring data indicated that the levels of PBDEs are increasing in aquatic organisms with higher concentr ations 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 usef ul to determine the extent of contamination in environmental media, and also the mechanisms of human exposure to this class of chemicals.PBBs and PBDEs 463

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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 in formation about average daily intake of PBBs was

located. The levels of PBBs in human tissue and bod y 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 Mi chigan to PBBs. This in formation 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. Th e 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 literatu re for blood, breast milk, and adipose tissue of the general population and occ upationally exposed individuals (WHO 1994a). Limited information on the levels of PBDEs in body tissues and fluids from indivi duals 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 brom inated BDEs in body fluids are a factor of 10– 100 higher for individuals living in the United States co mpared 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 partic ular, 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. Th e 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 studiesPBBs and PBDEs 464

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that examine these trends would be useful. Addition al 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 cond uct 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 pathwa y appears to be consumption of contaminated mother’s milk (see Section 8.4.4). More data ar e 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 recreati onal and subsistence fishers. Inform ation related to the exposure of children living near hazardous waste site s 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 lo cated in these areas. Quantitative information regarding the bioavailability and amount of PBBs th at children are exposed to through contact with contaminated soils are unavailable. Therefore, any in formation concerning this subject would be useful in evaluating children’s exposure. Additional informati on on weight-adjusted intakes would be helpful for determining the health risks for young children. In fants and young children consume a greater amount of food per kilogram of body weight and, therefore, ma y 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 cons umption 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 mamma ls. Information related to the exposure of children living near hazardous waste sites is also needed. In particular, in formation 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-PBBs and PBDEs 465

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to-mouth activity within homes located in these areas. Quantitative information regarding the bioavailability and amount of PBDEs that children ar e exposed to through contact with contaminated soils is unavailable. Therefore, any information c oncerning 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 tho se in Native American populations. Infants and young children consume a greater amount of food per kilogr am 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 Mich igan 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 samp les for PBB from 1975 to 1978. MDCH continues contact with this cohort, updates health questionnair es, 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 subregis try has been established in the National Exposure Registry. The substance will be considered in th e future when chemical selection is made for subregistries to be established. The information that is amassed in the National Exposure Registry facilitates the epidemiological resear ch 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.PBBs and PBDEs 466

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Polybrominated Diphenyl Ethers. A search of FEDRIP (2003) did not identify ongoing research studies for PBDEs. However, EPA has funded work in the Sc ience 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 st udies 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 bi phenyls (PBBs) and polybromintaed diphenyl ethers (PBDEs), its metabolites, and other biomarkers of expos ure and effect to PBBs and PBDEs. The intent is not to provide an exhaustive list of analytical me thods. Rather, the intention is to identify wellestablished methods that are used as the standard met hods of analysis. Many of the analytical methods used for environmental samples are the methods appr oved by federal agencies and organizations such as

EPA and the National Institute for O ccupational Safety and Health (NIO SH). 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 Associati on (APHA). Additionally, analytical methods are included that modify previously used methods to obta in lower detection limits and/or to improve accuracy and precision.

PBBs and PBDEs are analyzed in environmental and bi ological 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 year s between analytical efforts directed at PBBs and PBDEs, there have been many advances in the techno logy and costs of analytical instruments. Thus, while gas chromatography-electron capture detecti on (GC-ECD) with packed columns (i.e., noncongener specific) was the primary analytical technique u sed in the 1970s for PBBs, gas chromatography-mass spectrometry (GC-MS) with capillary columns (i.e., conge ner specific) is the primary analytical technique now used for PBDEs. These points should be consid ered 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 stor age, sample pretreatment, extraction, cleanup and fractionation, and analytical determ ination. Care must be taken to assure that the sample collection follows quality-assurance protocols and that equipment a nd containers are free from contamination. It is important that laboratories utilize blanks when reporti ng trace levels of PBBs and PBDEs. This practice will minimize the influence of trace contamination samp les that can originate from a variety of sources.PBBs and PBDEs 468

9. ANALYTICAL METHODS

Most sample collections are by grab sampling; how ever, 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 biol ogical 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 extr action techniques. Typical solvents are hexane, toluene, hexane/acetone mixtures, or dichloromethane. New extractions techniques, such as accelerated solvent extraction (ASE) or microwave-assisted extrac tion (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 so lid-phase trapping has been used for the extraction

brominated flame retardants from sediment with CO 2 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 an d 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., o ils 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, Flor isil) is used to remove other matrix interferences and to fractionate samples (Covaci et al. 2003).

The identification and quantitation of PBBs and PBDE s are most often accomplished by GC techniques. Capillary or high-resolution gas chromatography (H RGC) 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. Th e more universal and less sensitive flame-ionization detector (FID) is used much less often than the el ectron-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 identifica tion of PBB and PBDE congeners. A recent method of detection is electron-capture negative ionization (ECN I) as an ionization technique in combination withPBBs 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 mo re subject to brominated interferences and does not

allow the used of 13 C-labeled standards for quantification (Ikonomou and Rayne 2002). Conversely, electron ionization (EI) methods suffer from fragmenta tion 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 availa ble (e.g., 2,2,4,4’-tetrabromodiphenyl ether or BDE 47) for other members of its homolog group (e.g., tetrabro modiphenyl ether [tetraBDEs]). This can result in either under- or overestimating con centrations 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 degrad ation by UV light (i.e., bot h 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 contaminati on (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 accu rately determine the levels of BDE 209 and BDE 154 in analytical samples, analysts are required pe rform two separate GC measurements under different operating conditions.

9.1 BIOLOGICAL MATERIALS

Methods for the determination of organobromine co mpounds such as PBBs and PBDEs generally consist of the following steps: extraction of the analyte from the sample matrix; cleanup to remove interferingPBBs 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 de veloped for the determination of PBBs and PBDEs in blood or serum, urine, feces, adipose tissue, liver, a nd 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 w ith 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 extr act 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 lowe r 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 require d for the separation of the individual congeners in a mixture (Robertson et al. 1983b). However, decabro mobiphenyl 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 chromat ography, 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 μ g/kg in blood serum, 1 μ g/kg in human milk, and

0.5 μ g/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 ar e cleaned up to remove interferences using Florisil, silica gel, alumina or activated-charcoal column chro matography, gel-permeation chromatography (GPC),PBBs and PBDEs 471

9. ANALYTICAL METHODS



### Table 9-1. Analytical Methods for De termining 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 μ g/L 100.6–106.8 at 100 μ g/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 μ g/L (for hexa) 102 (for hexa) Willet et al. 1978

Whole blood Extract denatured sample with petroleum ether-ethyl ether; cleanup 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 μ g/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. 1983Liver 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 chromatogra phy; GPC = gel permeation chromatography; hexa = hexabrominated biphenyl; PBBs = polybrominated biphenyls; SI M = selected ion monitoringPBBs and PBDEs 472

9. ANALYTICAL METHODS



### Table 9-2. Analytical Methods for De termining 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 CapillaryGC-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 orLRGCHRMS (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 chromatograp hy; 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-t ert-butyl either; NCI = negative chemical ionization; PBDEs = polybrominated diphenyl ethers; SIM = sele cted ion monitoring; SPE = solid phase extractionPBBs 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. Hi gh recoveries (69–104%) of PBDE residues are obtained by the available analytical methods. Typically, th e 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 perform ed 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 fo r food; for PBDEs, detection limits are in the low ppb range for water matrices and in the low ppb to ppm ra nge for fish tissues. Analytical methods for the

determination of PBBs and PBDEs in environmen tal samples are given in Tables 9-3 and 9-4, respectively.

Polybrominated Biphenyls. Residues in environmental samples ca n be extracted using hexane-ether, petroleum ether-ether, toluene-ethyl acetate, or me thylene 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 (Robert son et al. 1983b). High recoveries (74–98%) of PBB residues in environmental samples are obtained by th e available analytical methods. Typically, the limit of quantitation for PBB residues is about 0.1 μ g/kg in soil and 0.7 μ g/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/h exane, 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 De termining 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 det ector; GC = gas chromatography; hexa = hexabrominated biphenyl; HRGC = high resolution gas chromatography; HRMS = high resolution mass spectrometry; LC = liquid chromatography; MS = mass spectrometry; NCI = negative c hemical ionization; PED = plasma emission detection; PBBs = polybrominated biphenylsPBBs and PBDEs 475

9. ANALYTICAL METHODS



### Table 9-4. Analytical Methods for De termining 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 μ g/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 μ g/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 GPCand 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 chro matography; GPC = gel permeation chromatography; HRMS = high resolution mass spectrometry; LC = liquid chromatography; MS = mass spectrometry; NAA = neutron activation analysis; NCI = negative chemical i onization; PBDEs = polybrominated diphenyl ethersPBBs 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 inte rferences using Florisil, silica gel, alumina or

activated charcoal column chro matography, gel permeation chroma tography (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. 1 997). 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 μ g/kg in sediment, and

0.2 μ g/kg lipid in fish (Akutsu et al. 2001; Yamamoto et al. 1997). The first inter-laboratory study on PBDEs in environmental samples showed that ther e is good agreement for quantification of BDE 47 and 2,2’,4,4’,6-pentabromodiphenyl ether (BDE 100) co ngeners. However, impr oved 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 ethe r (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 a nd 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 (a nd 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 substan ce-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 f illed. In the future, the id entified data needs will be evaluated and prioritized, and a substance-sp ecific 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 P BBs and PBDEs are available (Brilliant et al. 1978; Covaci et al. 2002b; Eyster et al. 1983; Land rigan 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 a nd 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 an d PBDEs analysis. Only 30–40 congener standards are currently available for identification and quantific ation of PBDEs (Eljarrat et al. 2002; Sjödin et al. 1998). Metabolites are also important biomarkers fo r 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; Jac obs et al. 1976, 1978; Yamamoto et al. 1997).

It would be helpful to develop da ta determining the detection limit a nd 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 lo wer brominated PBBs in environmental samples are available (Morris et al. 1992). An analytical me thod 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 brom inated products, which are the only environmental degradation products identified for PBBs. Anal ytical 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 lite rature of detectable biodegradation of PBBs in the environment under aerobic conditions (Griffin and Chou 1981a, 1981b), but the compounds mayPBBs and PBDEs 478

9. ANALYTICAL METHODS

biodegrade to debrominated products under anaerobic conditions in pollu ted environments (Morris et al. 1992).

It would be helpful to develop da ta determining the accuracy of PB DE determinations (e.g., percent recovery) in environmental samples. Methods for determining degradation pr oducts and metabolites of

PBDE are needed. There is no information in the litera ture 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 (PBD D/Fs), is often disturbed by the presence of PBDEs. Ebert et al. (1999) demonstrat ed that by using a Florisil column in a sample clean-up process, almost complete separation of PB DEs 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 acu te-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 derive d because serious developmental eff ects (fetal abortion and stillbirths) were observed in monkeys that had been exposed to P BBs 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 eff ects 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 ch ronic 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 T 4 hormone in fetal rats that were expo sed to a commercial pentaBDE mixture on days 4–20 of gestation (Zhou et al. 2002). A TSDR also derived an MRL of 0.007 mg/kg/day for intermediate-duration oral exposure to lower brominat ed 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 interm ediate-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/m 3 for intermediate-duration inhalation exposure to lower brominated BDEs. The intermediate inhalation MRL is based on a NOAEL of 1.1 mg/m 3 for thyroidPBBs and PBDEs 480

10. REGULATIONS AND ADVISORIES



### Table 10-1. Regulations and Guide lines Applicable to PBBs and PBDEs a

Agency Description Information Reference

INTERNATIONAL Guidelines: IARC Carcinogenicity classification IARC 1987 Decabromobiphenyl Group 2B b Decabromodiphenyl ether Group 3 c Hexabromobiphenyl Group 2B b Octabromobiphenyl Group 2B b

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 owne rs and operators of hazardous waste TSD facilities; groundwater monitoring for p- bromodiphenyl ether EPA 2002e

Suggested EPA method 8270 40CFR264, PQL 10 μ g/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 C d Nonabromodiphenyl ether Group D e Octabromodiphenyl ether Group D e Hexabromodiphenyl ether Group D e Pentabromodiphenyl ether Group D e Tetrabromodiphenyl ether Group D e Tribromodiphenyl ether Group D e p,p’- Dibromodiphenyl ether Group D e

p- Bromodiphenyl ether Group D e EPA Oral reference dose (RfD) IRIS 2002 Decabromodiphenyl ether 1x10 -2 mg/kg/dayPBBs and PBDEs 481

10. REGULATIONS AND ADVISORIES



### Table 10-1. Regulations and Guide lines Applicable to PBBs and PBDEs a

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 μ g/L

8270 10 μ g/L Radiation protection; health and environmental protection standards for uranium and thorium mill tailings; listed constituent EPA 2002d 40CFR192, Appendix I

p- Bromodiphenyl etherPBBs and PBDEs 482

10. REGULATIONS AND ADVISORIES



### Table 10-1. Regulations and Guide lines Applicable to PBBs and PBDEs a

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 InformedConsent 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 substancesknown 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 reporting f 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 06/30/98PBBs and PBDEs 483

10. REGULATIONS AND ADVISORIES



### Table 10-1. Regulations and Guide lines Applicable to PBBs and PBDEs a

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 μ g/L HSDB 2002

p- Bromodiphenyl ether c. Food No data d. Other

Florida Toxic substance BLR 2002 Polybrominated biphenylPBBs and PBDEs 484

10. REGULATIONS AND ADVISORIES



### Table 10-1. Regulations and Guide lines Applicable to PBBs and PBDEs a

Agency Description Information Reference

STATE (cont.) Massachusetts Hazardous substances requires the labeling of containers of toxic substances in the workplace BLR 2002

Decabromobiphenyl Extraordinary hazardous g Decabromdiphenyl ether Hexabromobiphenyl Extraordinary hazardous g Hexabromo-1,1'-biphenyl Extraordinary hazardous g Octabromobdiphenyl Extraordinary hazardous g

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 s ubstance; requires SICemployers 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 poundsPBBs and PBDEs 485

10. REGULATIONS AND ADVISORIES



### Table 10-1. Regulations and Guide lines Applicable to PBBs and PBDEs a

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: decabro mobiphenyl (CAS# 1 3654-09-6); 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); nonabromod iphenyl 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# 595 36-65-1); 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 s ubstances are subject to all provisions of 40CFR716. Manufacturers, importers, and processors of a listed subs tance are subject to the reporting requirements of Subpart A for that substance. g Extraordinary hazardous and designated carcinogen

ACGIH = American Conference of Governmental Industr ial 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 = Ha zardous 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 Toxicol ogy 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 averagePBBs 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) fo r 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 (2 002) has classified decabromobiphenyl, octabromobiphenyl, and polybrominated biphenyl as reasonably an ticipated to be human carcinogens. EPA (IRIS 2002) has classified decabromodiphenyl ether in Grou p D, as a possible human carcinogen, and nonabromodiphenyl ether, octabromodiphe nyl ether, hexabromodiphenyl ethe r, 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 (A IHA) has established a Workplace Environmental Exposure Level (WEEL) of 5 mg/m 3 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).PBBs and PBDEs 487



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### 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 molecu les (as of gases, solutes, or liquids) to the surfaces of solid bodies or liquids with which they are in contact.

Adsorption Coefficient (K oc )— The ratio of the amount of a chemical adsorbed per unit weight of organic carbon in the soil or sediment to the concentr ation 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 specifi ed adverse response. For example, a BMD 10 would be the dose at the 95% lower confidence limit on a 10% res ponse, and the benchmark response (BMR) would be 10%. The BMD is determined by modeling the dose response curve in the regi on 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, marker s 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 (o r 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 va riety of potential causative agents (such as toxic chemicals). In a case-controlled study, a group of pe ople 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 scien tific research, but are not actual research studies.PBBs and PBDEs 560

12. GLOSSARY

Ceiling Value— A concentration of a substance that shoul d 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 speci fic 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 leas t 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 ch emical 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 (e ither parent), during prenatal development, or postnatally to the time of sexual matu ration. Adverse developmental eff ects 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 variati ons, 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 effect s information. A health advisory is not a legally enforceable federal standard, but serves as technical guid ance 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 de fined 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, cl astogenic, 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 pop ulation 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 th e 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) (LC LO )— The lowest concentration of a chemical in air that has been reported to have caused death in humans or animals.

Lethal Concentration(50) (LC 50 )— 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) (LD LO )— The lowest dose of a chemical introdu ced by a route other than inhalation that has been reported to have caused death in humans or animals.

Lethal Dose(50) (LD 50 )— The dose of a chemical that has been calculated to cause death in 50% of a defined experimental animal population.

Lethal Time(50) (LT 50 )— 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 st atistically or biologically significan t 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 adve rsely 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 adver se noncancer health effects over a specified route and duration of exposure.

Modifying Factor (MF)— A value (greater than zero) that is appl ied to the derivation of a Minimal Risk Level (MRL) to reflect additional concerns about th e 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 numbe r 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 tissu es 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 (K ow )— 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 computationa l 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. Ther e 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, identif iable 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 rela tionship between target tissue dose and toxic end points. These models advance the importance of phys iologically 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 membra ne 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 pe rtinent 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 q 1 * can be used to calculate an estimate of carcinogenic potency, the incremental excess cancer risk per unit of exposure (usually μ g/L for water, mg/kg/day for food, and μ g/m 3 for air).

Recommended Exposure Limit (REL)— A National Institute for Occ upational Safety and Health (NIOSH) time-weighted average (TWA) concentrati ons 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 (inc luding 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 c ontinuous inhalation exposures and is appropriately expressed in units of mg/m 3 or ppm.

Reference Dose (RfD)— An estimate (with uncertainty spanning pe rhaps an order of magnitude) of the daily exposure of the human population to a potential hazar d that is likely to be without risk of deleterious effects during a lifetime. The RfD is operationally derived from the no-observe d-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, Compen sation, and Liability Act (CERCLA). Reportable quantities are (1) 1 pound or greater or (2) for select ed substances, an amount established by regulation either under CERCLA or under Section 311 of the Cl ean 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 di rected to the reproductive organs and/or the related endocrine system. The manifestation of such toxici ty may be noted as alterations in sexual behavior, fertility, pregnancy outcomes, or modifications in othe r functions that are dependent on the integrity of this system.

Retrospective Study— A type of cohort study based on a grou p 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 cau sal 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 a llowed per day, and there must be at least 60 minutes between exposure periods. The daily Threshold Limi t 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) ex tending 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 Gove rnmental 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 (R fC) 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 lifetim e 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, Compen sation, and Liability Act (CERCLA) [42 U.S.C. 9601 et seq.], as amended by the Superfund Amendmen ts 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 da ta needs associated with the substances.

The toxicological profiles include an examination, su mmary, and interpretation of available toxicological information and epidemiologic evaluations of a hazardous substance. During the development of toxicological profiles, Minimal Risk Levels (MRLs) ar e derived when reliable and sufficient data exist to identify the target organ(s) of effect or the most sen sitive 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 a dverse 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 contamin ants 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 a dverse health effects in the people most sensitive to

such chemical-induced effects. MRLs are derived fo r 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 gene rally based on the most sensitive chemical-induced end point considered to be of relevance to humans. Seri ous 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 ad verse 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 thatPBBs 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 th e 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 addre ss this uncertainty consistent with the public health principle of prevention. Although human data are pr eferred, MRLs often must be based on animal studies because relevant human studies are lacking. In th e absence of evidence to the contrary, ATSDR assumes

that humans are more sensitive to the effects of hazard ous 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 comment s 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 prev iously published levels. For additional information regarding MRLs, please contact the Division of To xicology, 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/m 3

Reference : Great Lakes Chemical Corporation. 2000. A 90-day inhalation toxicity study of octabromodiphenyl oxide in albino rats, dated 04/04/0 2. Submitted to the U.S. EPA under TCSA Section 8E, Fiche no. OTS0574171-1.

Experimental design : This is an unpublished industry-sponsor ed 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 fe male 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/m 3 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 T 3 , and total T 4 ) evaluations were performed near the end of the exposure period. Urinalyses were not conducted. Comprehensive necropies, organ weight measuremen ts, and histological examinations (including respiratory tract and thyroids) were perf ormed 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/m 3 and liver weight (absolut e and relative) at 202 mg/m 3 . Total incidences of centrilobular hepatocellular hype rtrophy in the 0, 1.1, 16, and 202 mg/m 3 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/m 3 , but showed no clear dose-related increasi ng 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/m 3 ; incidences in 0, 1.1, 16, and 202 mg/m 3 exposure groups were 4/10 (all minimal), 9/10 (7 minimal, 2 mild), 6/10 (all minima l), 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/m 3 (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/m 3 . Total incidences of alveolar histiocytosis at 0, 1.1, 16, and 202 mg/m 3 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/m 3 ; these included lung firmness and white discolo ration and/or enlargement in the bronchialPBBs and PBDEs A-4

APPENDIX A

and/or mediastinal lymph nodes. The lymph node e ffects 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 decr eases in mean thyroxine (total T 4 ) at ≥ 16 mg/m 3 in both sexes, and increases in thyroid stimulating hormone (TSH) at ≥ 16 mg/m 3 in males and 202 mg/m 3 in females. The changes were usually statistically si gnificant (p<0.05 or p<0.01) compared to controls and considered to be consistent with chemical-induced hy pothyroidism. There were no serum T 3 changes. Qualitative histological evaluations of step sections of ovaries show ed an absence of corpora lutea in 3/10 females at 202 mg/m 3 , compared to 0/10 in the control and lowe r 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/m 3 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, hemato crit, or hemoglobin levels). Serum chemistry evaluations showed that cholesterol was significantly increased (39.8% less than controls, p<0.01) in 202 mg/m 3 females, but the magnitude of the elevation w as not considered toxicologically significant. Some other statistically significant serum chemistr y alterations (increased mean globulin and total protein, decreased albumin/globulin ratio) also occurred in the 202 mg/m 3 females, but were not considered exposure-related due to sm all magnitudes of changes and lack of similar findings in the males.

Dose and end point used for MRL derivation : 1.1 mg/m 3

[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 th is nasal effect, identification of both a NOAEL (1.1 mg/m 3 ) and LOAEL (16 mg/m 3 ) 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 animal s 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 interm ittent to continuous exposure? The NOAEL was adjusted to continuous exposure as follows: 1.1 mg/m 3 x 6 hours/24 hours x 5 days/7 days = 0.196 mg/m 3

If an inhalation study in animals, list conversion f actors used in determining human equivalent dose : The human equivalent NOAEL (NOAEL HEC ) was calculated from the duration-adjusted NOAEL (NOAEL ADJ ) using EPA RfC methodology as follows:

NOAEL HEC = NOAEL ADJ x RDDR = 0.196 mg/m 3 x 2.7 = 0.53 mg/m 3

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 calculatePBBs and PBDEs A-5

APPENDIX A

the RDDR of 2.7: mean particle size (MMAD) of 2.0 μ m 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 = NOAEL HEC ÷ (UF x MF) = 0.53 ÷ (30 x 3) = 0.006 mg/m 3

Other additional studies or pertinent info rmation that lend support to this MRL : This is the only intermediate-duration inhalation study of PBDEs.

The thyroid is a sensitive target of lower brominat ed BDEs in orally exposed animals. A NOAEL for reduced serum T 4 hormone levels in fetal rats that were e xposed to pentaBDE (Zhou et al. 2002) was used as the basis for the acute oral MRL for lower brominat ed BDEs. This study is supported by findings of reduced serum T 4 levels in weanling rats that were acu tely exposed to octaBDE and pentaBDE as neonates (Zhou et al. 2001). Thyroid effects that mainly included reduced serum T 4 hormone levels and follicular cell hyperplasia were obser ved in a number of other animal studies of acute- or intermediateduration oral exposure to penta- a nd octaBDEs (Fowles et al. 1994; Hallgren et al. 2001; WIL Research Laboratories 1984).

Agency Contact (Chemical Manager) : Dr. Hana PohlPBBs and PBDEs A-6

APPENDIX A



### MINIMAL RISK LEVEL (MRL) WORKSHEET

Chemical name: Polybrominated Biphenyls (PBBs) CAS number(s): 36355-01-8 (unspeci fied 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 T 4 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 131 I, incorporation of 131 I into monoiodotyrosine, diiodotyrosine, T 3 or T 4 , amount of intrathyroidal iodide, thyroid and liver weights, and body weights. Differences between mean values for the measured parameters in the cont rol 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 (T 4 ) 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 131 I and a significant depression (p<0.01) in the incorporation of 131 I into monoiodotyrosine, without any a pparent effect on the incorporation of 131 I into diiodotyrosine, T 3 or T 4 . There was a significant increase (p<0.01) of intrathyroidal iodide (nine rats/dose evaluated). At ≥ 3 mg/kg, the absolute thyroid weights we re 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 observe d. 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 manuscrip t 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 variabilityPBBs 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 interm ittent to continuous exposure? NA

If an inhalation study in animals, list conversion f actors used in determining human equivalent dose : NA

Other additional studies or pertinent info rmation that lend support to this MRL : It is well documented in intermediate-duration studies that the thyroid is a targ et of PBBs showing a spectrum of effects, including decreases in serum T 3 and T 4 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 PohlPBBs 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/m 3

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 T 4 and T 3 ) concentrations, liver weight, and hepatic microsomal enzyme (EROD, PROD, UDPGT) activities. Serum T 3 was not assayed in GD 20 fetuses due to insufficient serum sample volum e. 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. An alysis of variance (ANOVA) was used to analyze for effects of treatment and interactions. If there w as 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 wei ghts (PNDs 4–21), and eye opening. Postweaning offspring body weights (PNDs 36–91) were analyz ed 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 T 4 data were analyzed with the Kruskal-Wallis test followed by a Dunn Multiple Co mparison tests (due to a lack of homogeneity of variance). A significance level of 0.05 was used for a ll 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 survivi ng until day 4), maternal or offspring body weights, or offspring eye opening were observed. The perina tal maternal exposure to PentaPBE caused significant (p<0.05) decreases in serum total T 4 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 T 4 concentrations in offspring was age-depende nt as values returned to control levels by PND 36. There were no exposure-related effects on serum total T 3 concentrations at any time in the damsPBBs and PBDEs A-9

APPENDIX A

or offspring. Relative liver weight was significan tly (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 cont rol 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 offs pring 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 T 4 (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 T 4 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 va riability 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 interm ittent to continuous exposure? NA

If an inhalation study in animals, list conversion f actors used in determining human equivalent dose : NA

Other additional studies or pertinent info rmation that lend support to this MRL : The NOAEL of 1 mg/kg/day and LOAEL of 10 mg/kg/day for reduced serum T 4 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 T 4 levels in weanling rats that were exposed to octaBDE for 4 days (Zhou et al. 2001). Thyroid hormone levels were determ ined 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 do ses 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 ev aluated for changes in serum total T 4 , total T 3 , and TSH, hepatic microsomal EROD, PROD, and UDPGT activities, and bod y 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-re lated reductions in serum total T 4 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 T 3 levels were reduced less than T 4 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 werePBBs and PBDEs A-10

APPENDIX A

no exposure-related changes in serum total TSH conc entrations. 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 (LED 20 ) were 9.25 and 5.29 mg/kg/day, respectively, for serum T 4 and 53.38 and 11.98 mg/kg/day, respectively, for serum T 3 . 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 T 4 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 T 3 levels were reduced less than T 4 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 cha nges in serum total TSH concentra tions. 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 (LED 20 ) were 12.74 and 6.95 mg/kg/day, respectively, for serum T 4 and 32.94 and 8.56, mg/kg/day, respectively, for serum T 3 . 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 T 4 hormone levels and follicular cell hyperplasiahave been observed in a number of other studies of PBDEs in orally-exposed animals. Other acuteduration studies showed decreases in serum T 4 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 T 4 in rats exposed to ≥ 10 mg/kg/day pentaBDE for 90 days (WIL Research Laboratories 1984). Intermediate-duration expo sure 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 hype rplasia developed in mice exposed to 2,240 mg/kg/day (NTP 1986).

Agency Contact (Chemical Manager) : Dr. Hana PohlPBBs 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/m 3

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 recove ry period (5 rats/sex/group) or 24-week recovery period (5 rats/sex/group). Animals were observed dail y 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 week s of the recovery period. Hematology, clinical chemistry (including serum T 3 and T 4 ), and urinalysis parameters, urin e and liver porphyrin levels, and serum bromide levels were evaluated in 10 rats/sex/g roup 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, ki dneys, 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-ta iled 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 abso lute 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. Hepatocy tomegaly 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 a nd 0/10, 0/10, 8/10, and 10/10 (statistical tests not conducted) in females. Some hepatocytes in aff ected areas had vacuoles, which were empty in tissue sections and reportedly likely containe d 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 gr oups appeared to have an increased incidence of degeneration and necrosis of i ndividual liver parenchymal cells at 24 weeks postexposure; thePBBs and PBDEs A-12

APPENDIX A

investigators concluded that this condition may repr esent the final loss of previously damaged cells and probably should be considered co mpound-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 hyperpl asia appeared to occur in lower do se 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 se verity 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 therefor e were mild and transient. Serum T 4 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 T 4 levels are likely indirect consequences of hepatic enzyme induction. No compound-related cha nges 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 porp hyrins 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 liv er 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 fro m 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 interm ittent to continuous exposure? NA

If an inhalation study in animals, list conversion f actors used in determining human equivalent dose : NA

Other additional studies or pertinent info rmation 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, li ver enlargement, and degenerative histopathologic alterations that progress to tumors. Repeated dietary exposure to PBDEs typically caused liver enlargement with or without degenerative changes, a nd 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 indu ced by chronic exposure to decaBDE included degeneration and thrombosis in ra ts 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 necr osis at higher doses) in rats exposed to octaBDE for 13 weeks (IRDC 1977), 8 mg/kg/day for hepatocellula r enlargement and vacuolation in rats exposedPBBs 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 PohlPBBs 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/m 3

Reference : Hardy ML, Schroeder R, Biesemeier J, et al. 2002. Prenatal oral (gavage) developmental toxicity study of decabromodiphenyl oxid e 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 impl antations, 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 fetu ses). The experiment was designed to meet current (as of 2002) EPA/OPPTS and OECD guidelines for a pr enatal 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 skelet al 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 de velopmental toxicity was found, i ndicating 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 interm ittent to continuous exposure? NA

If an inhalation study in animals, list conversion f actors used in determining human equivalent dose : NAPBBs and PBDEs A-15

APPENDIX A

Other additional studies or pertinent info rmation 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 (94

B 97% 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 9,50 0 mg/kg/day in mice. The daily doses (mg/kg body weight/day) were estimated by multiplying the dietar y 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 valu es for food consumption and body weight in F344 rats and B6C3F1 mice (E PA 1988a). All animals we re observed daily, weighed weekly, and necropsied at the end of the exposure peri od. 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 in termediate-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) st udy 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 PohlPBBs and PBDEs B-1



### APPENDIX B. USER'S GUIDE

Chapters 1 and 2

Public Health Statement

These chapters of the profile are health effect s summaries written in non-technical language. Their intended audience is the general public, especially peopl e living in the vicinity of a hazardous waste site or chemical release. If the Public Health Statem ents 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 H ealth Statements are useful to find specific topics of concern. The topics are written in a question and answer format. The answer to each questi on includes a sentence that will direct the reader to chapters in the profile th at 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, or al, and dermal) and within route by effect. Human data are presented first, then animal data. Both ar e organized by duration (acute, intermediate, chronic). In vitro data and data from parenteral routes (intramusc ular, 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. Minima l 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 preven t a satisfactory evaluation of the relevance to public health are identified in the Ch apter 5 Data Needs section.

Interpretation of Minimal Risk Levels

Where sufficient toxicologic information is availabl e, 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 he alth professionals with e xposure 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 pu blic health officials determine the safety of a community living near a chemical emission, given the concentration of a contam inant 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 info rmation on which the number is based. Chapters 3 and 4, "Relevance to Public Health," contain basi c information known about the substance. Other sections such as Chapter 5 Section 5.9, "Interac tions with Other Substances,” and Section 5.10, "Populations that are Unusually Susceptible" pr ovide important supplemental information.

MRL users should also understand the MRL deri vation methodology. MRLs are derived using a modified version of the risk assessment methodolog y 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 developm ental effects. If this information and reliable quantitative data on the chosen end point are availabl e, ATSDR derives an MRL using the most sensitive species (when information from multiple species is availa ble) with the highest no-observed-adverse-effect level (NOAEL) that does not exceed any adverse eff ect levels. When a NOAEL is not available, a lowest-observed-adverse-effect level (LOAEL) can be u sed to derive an MRL, and an uncertainty factor (UF) of 10 must be employed. Additional uncerta inty factors of 10 must be used both for human variability to protect sensitive subpopulations (peopl e who are most susceptible to the health effects caused by the substance) and for interspecies vari ability (extrapolation from animals to humans). In deriving an MRL, these individual uncertainty fact ors 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 effe cts and illustrate graphica lly levels of exposure associated with those effects. These levels cover health effects observed at increasing dose concentrations and durations, diffe rences 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 appli cation 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 princi pal routes of exposure, i.e., inhalation, oral, and dermal (LSE Table 5-1, 5-2, and 5-3, respec tively). LSE figures are limited to the inhalation (LSE Figure 5-1) and oral (LSE Figure 5-2) rout es. 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 prese nted within each relevant route of exposure. In this example, an inhalation study of interm ediate exposure duration is reported. For quick reference to health effects occurring from a know n 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 "S ystem" 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 correspondi ng 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 re levance of animal data to human toxicity and Section 5.4, "Toxicokinetics," contains any availa ble 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 an d 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), ra ts 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 th e original reference paper (i.e., Nitschke et al. 1981).

(7) System . This column further defines the systemic effects. These systems include respiratory, cardiovascular, gastrointestinal, hematologi cal, 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 numbe r 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" an d "Serious" effects. These distinctions help readers identify the levels of exposure at whic h adverse health effects first appear and the gradation of effects with increasing dose. A brie f 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 associ ated 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 refere nce 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 interm ediate exposure periods are illustrated.

(14) Health Effect . These are the categories of health effects for which reliable quantitative data exists. The same health eff ects 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. Exposur e concentration or dose is measured on the log scale "y" axis. Inhalation exposure is reported in mg/m 3 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 inhalati on 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 (q 1 *).

(19) Key to LSE Figure . The Key explains the abbreviations and symbols used in the figure.PBBs and PBDEs B-6

APPENDIX B10Reference↓Nitschke et al.1981Wong et al. 1982NTP 1982NTP 1982Serious (ppm)(CEL, multiple organs)(CEL, lung tumors, nasaltumors)(CEL, lung tumors,hemangiosarcomas)11↓201010LOAEL (effect)Less serious (ppm)9↓10 (hyperplasia)SAMPLENOAEL(ppm)8↓3b7System↓Resp6Exposurefrequency/duration↓13 wk5 d/wk6 hr/d18 mo5 d/wk7 hr/d89-104 wk5 d/wk6 hr/d79-103 wk5 d/wk6 hr/d5Species↓RatRatRatMouseTABLE 5-1. Levels of Significant Exposure to [Chemical x] - InhalationKey tofigureaINTERMEDIATE EXPOSURESystemic18CHRONIC EXPOSURECancer383940aThe number corresponds to entries in Figure 5-1.bUsed to derive an intermediate inhalation Minimal Risk Level (MRL) of 5x10-3ppm; dose adjusted for intermittent exposure and dividedby an uncertainty factor of 100 (10 for extrapolation from animal to humans, 10 for human variability).→→→→→123412PBBs and PBDEs B-7

APPENDIX BPBBs and PBDEs C-1



### APPENDIX C. ACRONYMS, AB BREVIATIONS, AND SYMBOLS

ACGIH American Conference of Gove rnmental Industrial Hygienists ACOEM American College of Occupa tional 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 Occupationa l 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 Res ponse, 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 TransportationPBBs and PBDEs C-2

APPENDIX C

DOT/UN/ Department of Transportation/United Nations/ NA/IMCO North America/Inte rnational Maritime Da ngerous 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 F 1 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, F ungicide, 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 K oc organic carbon partition coefficient K ow octanol-water partition coefficient L liter LC liquid chromatography LC 50 lethal concentration, 50% kill LC Lo lethal concentration, low LD 50 lethal dose, 50% kill LD Lo lethal dose, low LDH lactic dehydrogenase LH luteinizing hormone LOAEL lowest-observed-adverse-effect level LSE Levels of Significant Exposure LT 50 lethal time, 50% kill m meter MA trans,trans -muconic acid MAL maximum allowable level mCi millicurie MCL maximum contaminant levelPBBs 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 O ccupational 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 WaterPBBs and PBDEs C-4

APPENDIX C

OWRS Office of Water Regula tions 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 TD 50 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 OrganizationPBBs 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 μ m micrometer μ g microgram q 1 * cancer slope factor – negative + positive (+) weakly positive result (–) weakly negative resultPBBs 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, 7 5, 185, 210, 248, 250, 251, 253 adsorbed............................................................................................. 379, 382, 395, 396, 397, 3 98, 400, 401, 404, 406 adsorption .................................................................................................................. 357 , 395, 397, 398, 399, 400, 460 aerobic ....................................................................................................................... 4 03, 406, 407, 408, 460, 477, 478 Agency for Toxic Substances and Disease Re gistry................................................ 8, 9, 19, 20, 45, 247, 270, 288 , 317 Ah rece ptor .................................................................................................................... ........................................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.................................................................................................................... 38 1, 403, 406, 407, 408, 460, 478 anemia ......................................................................................................................... ......................................171, 322 antiestro genic................................................................................................................. ....................282, 286, 288, 301 Arctic ......................................................................................... 397, 410, 419, 430, 431, 432, 43 8, 439, 440, 453, 456 ATSDR.................................................................................. 8, 9, 19, 20, 35, 36, 60, 61, 308, 316, 317, 457, 476, 479 Bioaccumulation................................................................................................................ ................................400, 461 Bioavaila bility

................................................................................................................ ...........................................460 bioconcentrat ion factor ........................................................................................................ ..............................396, 399 Biodegradation ................................................................................................................. .................................382, 407 Biomagnification ............................................................................................................... ........................................400 biomarkers ......................................................................................................... 293, 303, 30 7, 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 Carcinog enic................................................................................................................... ...........................................210 Carcinog enic ity................................................................................................................ ............................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, 2 99, 304, 309, 338 computer monitors.............................................................................................................. .............................1, 14, 375 computer-dis assembly ........................................................................................................... ....................................447 congener ........................14, 24, 33, 40, 43 , 47, 57, 59, 60, 75 , 185, 190, 198, 202, 204, 20 8, 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, 1 93, 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, 461PBBs and PBDEs D-2

APPENDIX D

deoxyribonucl eic aci d.......................................................................................................... ......................237, 238, 240 Dermal Effects................................................................................................................. ............................76, 190, 231 Distributio n................................................................................................................ 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 detec tion..................................................................................................... .............471, 472, 474, 475 electron-capture ne gative ionization........................................................................................... ...............................468 electronic en closures.......................................................................................................... ..................................12, 374 Elimination .................................................................................................................... ....................248, 254, 261, 269 elimination half-time .......................................................................................................... .......................................254 elimination half-times......................................................................................................... .......................................254 elimination rate............................................................................................................... ...........................................304 endocrine ....................................................................... 31, 75, 18 3, 231, 275, 285, 286, 287, 288, 301, 322, 341, 343 Environmental Protec tion 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 ti ssue................................................................................................................... .............................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, 17 1, 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-ionizati on detector...................................................................................................... .....................................468 flexible polyuretha ne foam (FPUF).............................................................................................. .....................372, 384 Food and Drug Ad ministration................................................................................................... .......8, 19, 58, 216, 485 FSH............................................................................................................................ ........................................286, 295 gas chromatography........................................................................................... 259, 404, 467, 468 , 471, 472, 474, 475 Gastrointestin al Effects....................................................................................................... .................................71, 169 gel-permeation chromatography.................................................................................................. ..............................470 general population ...........................3, 8, 14, 18, 24, 25, 37, 39, 59, 292, 302, 303, 306, 312, 3 13, 315, 334, 381, 382

, 420, 450, 451, 452, 453, 456, 459, 462, 463, 464 genotoxic ...................................................................................................................... .................40, 60, 237, 241, 326 Genotoxi city ................................................................................................................... ...........................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, 34 0, 341, 354, 366, 368, 373, 374, 398, 410, 412, 417, 418, 420, 427, 441, 458, 486 groundwater......................................................................... 13, 38 1, 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 poly styrene (HIPS) ................................................................................................. ...............373, 374, 376 homolog........................................................................................................................ .............247, 345, 354, 361, 469 humic acid ..................................................................................................................... ....................................404, 468 hydroxyl radical............................................................................................................... ..........................................401 IARC ........................................................................... 6, 16, 33, 45, 353, 35 5, 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 ..................................................................................................................... ..............................................452PBBs and PBDEs D-3

APPENDIX D

Integrated Risk In formation System............................................................................................. .............................485 IRIS ........................................................................................................................... ................................480, 485, 486 IUPAC.......................................................................................................................... .....................346, 347, 354, 359 Kow ............................................................................................................................ ...............357, 361, 362, 399, 400 landfills.............................................................................. 37 5, 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 Corporatio n ............................................... 255, 346, 367, 374, 387, 392, 394, 396, 412, 413, 4 17 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 Musculoskeleta l Effects........................................................................................................ .....................................173 National Human Adipose Tissue Survey........................................................................................... ........................447 National Prio rities List....................................................................................................... ......................1, 11, 379, 383 National Toxicology Program .................................................................................................5, 33, 224, 317, 466, 485 neonatal ................................................................... 41, 42 , 56, 202, 204, 207, 2 41, 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, 31 6, 325, 330, 331, 332, 340, 343 neurodevelopmenta l......................................................................... 25, 40, 42, 201, 215, 297, 298, 299, 313, 331, 333 neurotrans mitter............................................................................................................... ..........................................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 coe fficients......................................................................................................... .........................................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 Photol ysis ..................................................................................................................... .............................401, 408, 459 placenta.................................................................................. 4, 7, 15, 17, 247, 24 9, 292, 302, 306 , 333, 336, 421, 463 placental tissue............................................................................................................... ............................................455 polybrominated dibenzofurans ................................................................................................... ...............358, 365, 375PBBs and PBDEs D-4

APPENDIX D

polybrominated-p-dibenzodioxins................................................................................................ .............................365 polyvinyl chloride (PVC) ....................................................................................................... ...................................373 publicly owned treatme nt works (POTWs) ......................................................................................... ..............375, 458 Pyrolysis ...................................................................................................................... ......................................358, 387 rate constant.................................................................................................................. .....................................269, 304 Regulations.................................................................................................................... ..................8, 19, 480, 483, 485 Renal Ef fects .................................................................................................................. .....................................73, 180 retention..................................................................................... 33, 39, 46, 60, 201, 251, 252, 25 8, 264, 299, 309, 470 RfD ............................................................................................................................ ................................480, 485, 486 rigid poly urethane............................................................................................................. .........................................373 sediments ............................................................................. 13, 341, 392, 397, 399, 403, 408, 413, 41 5, 416, 432, 462 sewage sludges ................................................................................................................. .........................................375 SGOT........................................................................................................................... ........................26, 174, 177, 310 solub ility .................................................................................................................... 2 44, 251, 258, 264, 364, 398, 400 SRBC........................................................................................................................... ......................196, 197, 198, 278 Superf und ...................................................................................................................... ....................................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, 2 14, 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 toxicoki netic .................................................................................................................. 39, 57, 299, 309, 313, 337, 338 toxicoki netics................................................................................................................. ......................46, 241, 297, 337 triiodothyr onine ...........................................................................................................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, 36 1, 363, 364, 384, 387, 389, 396, 397, 398, 453, 4 57, 463 vola tility

..................................................................................................................... ........389, 395, 397, 398, 400, 470 volatilization .............................................................................................. 323, 384, 387, 389, 395, 396, 397, 398, 459 water solubility .......................................................................................... 355, 357, 36 1, 363, 3 85, 396, 397, 398, 400 weanling .......................................................................................... 47, 52, 54 , 175, 179, 186, 18 7, 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, 3 99, 421, 453, 455, 456, 463, 464, 472, 473, 475, 476 World Trad e Center............................................................................................................. ................................13, 425