&EPA
United States
Environmental Protection Agency
EPA DocumentW 740-R1-4004
August 2015
Office of Chemical Safety and
Pollution Prevention
TSCA Work Plan Chemical
Problem Formulation and Initial Assessment
Tetrabromobisphenol A and Related Chemicals Cluster
Flame Retardants
Br
CASRN
79-94-7
21850-44-2
25327-89-3
37853-61-5
NAME
Phenol, 4,4'-(l-methylethylidene)bis[2,6-
dibromo
Benzene, l,l'-(l-methylethylidene)bis[3,5-
dibromo-4-(2,3-dibromopropoxy)-
Benzene, l,l'-(l-methylethylidene)bis[3,5-
dibromo-4-(2-propen-l-yloxy)-
Benzene, l,l'-(l-methylethylidene)bis[3,5-
dibromo-4-methoxy-
August 2015
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Table of Contents
List of Tables 4
List of Figures 4
AUTHORS / CONTRIBUTORS / ACKNOWLEDGEMENTS 5
ABBREVIATIONS 6
EXECUTIVE SUMMARY 9
1 INTRODUCTION 12
1.1 SCOPE OF THE ASSESSMENT 13
1.2 REGULATORY AND ASSESSMENT HISTORY 14
1.2.1 Federal 14
1.2.2 State 14
1.2.3 International 15
1.2.3.1 European Union 15
1.2.3.1 Canada 16
2 PROBLEM FORMULATION 18
2.1 PHYSICAL AND CHEMICAL PROPERTIES 19
2.2 PRODUCTION VOLUME AND USES 20
2.2.1.1 Production 20
2.2.1.2 Uses 22
2.2.1.3 Summary of Production and Use 26
2.3 FATE AND TRANSPORT 26
2.4 EXPOSURES 27
2.4.1 Releases to the Environment 27
2.4.1.1 Chemical Manufacturing 27
2.4.1.2 Processing 27
2.4.1.3 Recycling 27
2.4.1.4 Disposal 28
2.4.2 Presence in the Environment and Biomonitoring Data 29
2.4.3 Occupational Exposures 30
2.4.4 General Population Exposures 30
2.4.5 Consumer Exposures 31
2.5 HAZARD ENDPOINTS 31
2.5.1 Ecological Hazard 31
2.5.2 Human Health Hazard 32
2.6 RESULTS OF PROBLEM FORMULATION 32
2.6.1 Conceptual Models 32
2.6.2 Analysis Plan 40
2.6.2.1 Environmental Assessment near Manufacturing Facilities 40
2.6.2.2 Human Health Risk Assessment 41
2.6.3 Sources and Pathways Excluded from Further Assessment 42
2.6.3.1 Chemical Manufacturing 42
2.6.3.2 Processing 43
2.6.3.3 Recycling 43
2.6.3.4 Disposal 44
2.6.3.5 Other Excluded Pathways 45
2.6.4 Uncertainties and Data Gaps 46
2.6.4.1 Environmental Fate Data 46
2.6.4.2 Release Data 46
2.6.4.3 Exposure Information 46
2.6.4.4 Ecological Hazard Data 47
2.6.4.5 Human Health Hazard Data 47
References 49
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Appendix A Data Available for TBBPA-bis(dibromopropyl ether), TBBPA-bis(allyl ether) and TBBPA-
bis(methyl ether) 71
A-l EXCLUSION FROM FURTHER ASSESSMENT 73
A-2 CHEMICAL STRUCTURES 74
A-3 PHYSICAL-CHEMICAL PROPERTIES 75
A-4 PRODUCTION VOLUMES 76
A-5 USES 77
A-6 FATE PROPERTIES 80
A-7 EXPOSURE 81
A-8 ECOLOGICAL HAZARD 83
A-9 HUMAN HEALTH HAZARD 84
Appendix B Literature Searches and Data Quality 86
B-l LITERATURE AVAILABLE FOR THE CLUSTER MEMBERS 86
B-2 DATA ADEQUACY 87
Appendix C Fate and Transport 88
C-l FATE IN ENVIRONMENTAL MEDIA 89
C-2 PERSISTENCE 90
C-3 BlOACCUMULATION/BlOCONCENTRATION 91
C-4 EXCLUSION OF DEGRADATION PRODUCTS FROM FURTHER ASSESSMENT 91
Appendix D Toxics Release Inventory Emissions 93
D-l TRI RELEASES FROM MANUFACTURERS 93
Appendix E TBBPA and other Dust Particle Concentrations in Occupational Settings 97
E-l PROCESSING 98
Appendix F Ecological Hazard Study Summaries 99
F-l TOXICITY TO AQUATIC ORGANISMS 99
F-2 TOXICITY TO TERRESTRIAL ORGANISMS 105
F-3 TOXICITY TO AMPHIBIANS 107
F-4 SUMMARY OF ENVIRONMENTAL HAZARD 108
Appendix G Human Health Hazard Study Summaries 109
G-l EPIDEMIOLOGY 109
G-2 TOXICOKINETICS 109
G-3 ACUTE TOXICITY 110
G-4 REPEATED-DOSE TOXICITY 110
G-5 REPRODUCTIVE AND DEVELOPMENTAL TOXICITY 112
G-6 IRRITATION AND SENSITIZATION 115
G-7 GENOTOXICITYANDCARCINOGENICITY 115
G-8 STUDIES PROPOSED FOR RISK ASSESSMENT 122
G-9 UNCERTAINTIES 123
Appendix H Parameters Needed for Estimating Fish Ingestion 124
Appendix 1 Draft Approach for Estimating Exposure from Mouthing of TBBPA 126
1-1 TBBPA CONCENTRATIONS IN PRODUCTS AND PRODUCT SURFACES 126
1-2 MIGRATION RATES INTO SALIVA 127
1-3 CHILDREN'S ACTIVITY PATTERNS 127
1-4 OTHER PARAMETERS BEING CONSIDERED 128
1-5 METHOD TO COMBINE EXPOSURE DATA 128
1-6 CONFIDENCE AND UNCERTAINTY IN THE AVAILABLE DATA 128
Appendix J Dose-Response Assessment for Cancer Endpoints 129
J-l CHOICE OF MODEL, POINTS OF DEPARTURE AND ORAL SLOPE FACTORS 129
J-2 CALCULATION OF TARGET RISK LEVELS 130
J-3 KEY SOURCES OF UNCERTAINTY 131
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Appendix K EU and Canada Risk Assessments: Specific Evaluations Consulted for Current
Assessment 133
K-l EU CALCULATIONS OF TBBPA UPTAKE FROM SOIL NEAR MANUFACTURING AND PROCESSING SITES 133
K-2 EU ENVIRONMENTAL ASSESSMENT AT PROCESSING SITES 133
List of Tables
Table 1-1: Chemical Names and Structures 13
Table 2-1: Data Required for Risk Assessment 19
Table 2-2: Physical-Chemical Properties 20
Table 2-3: CDR Production Volume Data 21
Table 2-4: Industrial and Consumer Use Data for TBBPA from the CDR 23
Table 2-5: Availability of Exposure Data for TBBPA 29
Table 2-6: Environmental Exposure Scenarios Considered for Assessment 35
Table 2-7: Human Health Exposure Scenarios Considered for Assessment 38
Table_Apx A-l: Data Availability and Read Across for Cluster Members 71
Table_Apx A-2: Physical-Chemical Properties 76
Table_Apx A-3: 2012 CDR Production Volume Data (Pounds/Year) for TBBPA-bis(dibromopropyl ether), TBBPA-bis(allyl
ether) and TBBPA-bis(methyl ether) 77
Table_Apx A-4: Industrial and Consumer Use Data for TBBPA-bis(dibromopropyl ether), TBBPA-bis(allyl ether) and TBBPA-
bis(methyl ether) 79
Table_Apx A-5: Environmental Fate Endpoints for Three Cluster Members 80
Table_Apx A-6: Availability of Exposure Data for Three Cluster Members 82
Table_Apx C-l: Environmental Fate 88
Table_Apx D-l: Disposal and Releases of TBBPA by Industry as Reported in the 2012 TRI 93
Table_Apx D-2: Disposal and Releases of TBBPA by Industry as reported in the 2012 TRI 94
Table_Apx D-3: Disposal and Releases of TBBPA by Manufacturing Facility as Reported in the 2012 TRI 95
Table_Apx D-4: TBBPA TRI Release Trends from 2001 to 2012 95
Table_Apx D-5: Air Emissions for Facilities Modeled in the Current Assessment 96
Table_Apx E-l: Potential Occupational Exposures from Manufacture of TBBPA* 97
Table_Apx E-2: Potential exposures from processing of TBBPA* 98
Table_Apx F-l: Toxicity from TBBPA to Aquatic Organisms 102
Table_Apx F-2:Soil Invertebrate Toxicity Data for TBBPA (mg/kg) 106
Table_Apx F-3: Ecotoxicity Concentrations of Concern 108
Table_Apx 1-1: Plan for Evaluating Risk to Children from Ingestion of TBBPA from Products in the Home 126
Table_ApxJ-l: Parameters Used in Dose-Response Equations 130
Table_Apx J-2: Doses (in mg/kg-bw/day) Associated with Three Target Risk Levels 131
List of Figures
Figure 1-1: General Structure for TBBPA and Related Chemicals 13
Figure 2-1: Conceptual Model for the TBBPA Environmental Assessment 34
Figure 2-2: Conceptual Model for the TBBPA Human Health Assessment 37
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AUTHORS / CONTRIBUTORS / ACKNOWLEDGEMENTS
This report was developed by the United States Environmental Protection Agency (US EPA), Office of
Chemical Safety and Pollution Prevention (OCSPP), Office of Pollution Prevention and
Toxics (OPPT). The Work Plan Chemical Problem Formulation for the cyclic aliphatic bromides cluster
was prepared based on currently available data. Mention of trade names does not constitute
endorsement by EPA.
EPA Assessment Team
Lead: Amy Benson, OPPT/Risk Assessment Division (RAD)
Team Members:
Charles Bevington, OPPT/RAD
Pamela Buster, OPPT/Environmental Assistance Division (EAD)
Rehan Choudhary, OPPT/RAD
Ana Corado, OPPT/EAD
Conrad Flessner, OPPT/RAD
Greg Fritz, OPPT/Chemistry, Economics and Sustainable Strategies Division (CESSD)
Jeff Gallagher, OPPT/RAD
Kirsten Hesla, OPPT/Chemical Control Division (CCD)
David Lai, OPPT/RAD
David Lynch, OPPT/RAD
Tim Lehman, OPPT/CESSD
Larry Newsome, OPPT/RAD
Eva Wong, OPPT/RAD
Yintak Woo, OPPT/RAD
Management Leads:
Stan Barone, OPPT/RAD
MarkTownsend, OPPT/RAD
Acknowledgements
The following individuals contributed to portions of this document:
Vivian Hart, OPPT/RAD
Ruth Hummel, OPPT/RAD
Emma Lavoie, OPPT/CESSD
Yvette Selby-Mohamadu, OPPT/RAD
Portions of this document were developed with support from Abt Associates, ICF International, Versar
and SRC, Inc.
Docket
Please visit the public docket (Docket: EPA-HQ-OPPT-2014-0730) to view supporting information.
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ABBREVIATIONS
ABS
ADEQ
AERMIC
AERMOD
AMS
ASTM
Bl, 2 or 3
B6C3F1
B6C3F1/N
BA-59P
BAEP
BAF
BCF
BPA
BSEF
CASRN
CBI
CFR
CDR
COC
DL
Doverguard 59
Doverguard 68
dw
EC
ECB
EC/HC
ECHA
ECOTOX
EPA
EPI
ER
EU
e-waste
Fo
Fi
F344/NTac
FR-720
FR-1524
GLCC
GLP
Acrylonitrile butadiene styrene
Arkansas Department of Environmental Quality
AMS/EPA Regulatory Model Improvement Committee
AERMIC AMS/EPA Regulatory Model
American Meteorological Society
American Society for Testing and Materials
Bioaccumulation level 1, 2 or 3
Mouse strain
Mouse strain
Trade name for TBBPA
Brainstem auditory evoked potential
Bioaccumulation factor
Bioconcentration factor
Bisphenol A
Brominated Science and Environmental Forum
Chemical Abstract Service Registry Number
Confidential business information
Code of Federal Regulations
Chemical data reporting
Concentration of concern
Detection limit
Trade name for TBBPA
Trade name for TBBPA-bis(dibromopropyl ether)
Dry weight
European Commission
European Chemicals Bureau
Environment Canada and Health Canada
European Chemicals Agency
EPA database of ecotoxicity studies
United States Environmental Protection Agency
Estimation Programs Interface
Estrogen receptor
European Union
electronic waste
Parental generation
First generation of offspring
Rat strain
Trade name for TBBPA-bis(dibromopropyl ether)
Trade name for TBBPA
Great Lakes Chemical Corporation
Good laboratory practice
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HPV
ICL-IP
ISO/DIS
IUR
kg
Ib
LC50
LD50
LOAEL
LOEC
LOQ
Log Koc
L°g Kow
Iw
MATC
mg/kg-bw/day
MITI
MOA
MOE
MOS
MSW
NAICS
NMRI
NOAEL
NOEC
NTP
OCSPP
OECD
OPPT
PI, 2 or 3
PAHs
PBDDs
PBDFs
PND
QSAR
RCRA
REACH
SAYTEX® CP-2000
T3
T4
TBBPA
TOC
Tp53
TRI
TSCA
High production volume
ICL Industrial Products America, Inc.
International Organization for Standardization/Draft International Standard
Inventory update reporting
Kilogram
Pound
Lethal concentration at which 50% of test organisms die
Lethal dose at which 50% of test organisms die
Lowest-observed-adverse-effect level
Lowest-observed-effect concentration
Limit of quantitation
Logarithmic soil organic carbon partition coefficient
Logarithmic octanol-water partition coefficient
Lipid weight
Maximum acceptable toxicant concentration
Milligrams (of a chemical) per kilogram of body weight per day
Ministry of International Trade and Industry (Japan)
Mode of action
Margin of exposure
Margin of safety
Municipal solid waste
North American Industry Classification System
Mouse strain
No-observed-adverse-effect level
No-observed-effect concentration
National Toxicology Program
Office of Chemical Safety and Pollution Prevention
Organisation for Economic Co-operation and Development
Office of Pollution Prevention and Toxics
Persistence level 1, 2 or 3
Polyaromatic hydrocarbons
Polybrominated dibenzodioxins
Polybrominated dibenzofurans
Postnatal day
Quantitative structure activity relationship
Resource Conservation and Recovery Act
Registration, Evaluation, Authorisation and Restriction of Chemicals
Trade name for TBBPA
Triiodothyronine
Thyroxine
Tetrabromobisphenol A
Total organic carbon
Gene for tumor suppressor p53 protein
Toxics Release Inventory
Toxic Substances Control Act
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TSH Thyroid stimulating hormone
TTR Transthyretin
UL Underwriters Laboratories
US United States
ww Wet weight
WWTP Wastewater treatment plant
yr Year
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EXECUTIVE SUMMARY
As part of EPA's comprehensive approach to enhance the Agency's management of existing chemicals,
EPA/OPPT identified a work plan of chemicals for further assessment under the Toxic Substances
Control Act (TSCA) in March 2012. Chemical risk assessments will be conducted if, as a result of scoping
and problem formulation, there are exposures of concern, identified hazards and sufficient data for
quantitative analysis. If an assessment identifies unreasonable risks to humans or the environment,
EPA will pursue risk reduction. This document presents the problem formulation for TBBPA and related
chemicals as part of the TSCA Work Plan.
EPA/OPPT included TBBPA-bis(dibromopropyl ether) (CASRN 21850-44-2) and TBBPA-bis(allyl ether)
(CASRN 25327-89-3) in the TBBPA cluster as a result of an initial prioritization exercise because these
compounds have additive flame retardant uses; EPA/OPPT assumes that additive uses will lead to
higher potential for exposure. EPA/OPPT included a fourth chemical, TBBPA-bis(methyl ether) (CASRN
37853-61-5), because monitoring studies have routinely found this substance in the environment. Its
presence is likely to be a result of microbial transformation of TBBPA.
Conclusions
As a conclusion of this problem formulation and initial assessment, EPA/OPPT will further assess the
following risks to:
Environment
• Aquatic, sediment-dwelling or soil-dwelling organisms resulting from two manufacturing
facilities that emitted the vast majority of TBBPA to air during a 13-year period (2000-2012) as
indicated from Toxics Release Inventory (TRI) data.
Human Health
• Workers at manufacturing and processing facilities who may ingest TBBPA in dust from the air
after further developing assessment methods.
• Aggregate oral exposure from the following oral exposure pathways:
o Incidental ingestion of TBBPA in dust from outdoor sources
o Incidental ingestion of TBBPA in dust from indoor sources
o Incidental ingestion of TBBPA from mouthing of consumer products
o Consumption of TBBPA in fish (recreational and subsistence fishers)
These aggregate exposures will be assessed and compared for the following: 1) those who live near
two manufacturing facilities, and 2) those who do not live near such facilities.
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EPA/OPPT will not specifically assess the following risks:
• To the general population near processing sites, due to smaller releases from processors than
from manufacturers.
• To the environment and human health from recycling plants given limited information
regarding recycling of electronic products in the United States.
• From product disposal, due to required controls, limited data or low concerns.
• From drinking water due to lack of information as well as a likelihood of low risk concerns.
• From eating food other than fish because this is considered the purview of other agencies.
• For adults exposed directly (through mouthing) to consumer products, based on low concerns.
Exclusion of Three Cluster Members and Degradation Products
Some limited information is available for the cluster members other than TBBPA. However, EPA/OPPT
concluded that no quantitative risk assessment is needed for these other cluster members for one or
more of the following reasons: limited information, inability to use the more robust data for TBBPA to
read across to other cluster members, low toxicity or likely low risk concerns.
EPA/OPPT also investigated the possibility of evaluating risks for degradation and combustion products
of TBBPA. However, information is too limited on rates of degradation and the specificity of identified
combustion products to incineration of TBBPA. Therefore, EPA/OPPT will not assess risks from
exposure to these products in the proposed assessment.
TBBPA Data Evaluated During Problem Formulation
TBBPA (CASRN 79-94-7) is used as a flame retardant, is persistent and has been detected in the
environment, in humans and in biota. Of the brominated flame retardants, TBBPA has the highest US
and global production volume with a reported 2011 US volume of 120 million pounds. It is used as both
an additive and reactive flame retardant1 and is reacted to produce more than 70 compounds. It is
used primarily in electronics but may be found in a variety of other products.
In 2012, reports to TRI indicated that 52 manufacturing and processing facilities released or disposed
approximately 127,845 pounds of TBBPA. The compound may undergo direct photolysis and indirect
photo-oxidation, but the amount of vapor in air is expected to be low. It is expected to be persistent in
water, soil and sediment and has a low bioaccumulation potential.
TBBPA has been found in humans (blood, breast milk and adipose tissue) and in biota (aquatic and
terrestrial animals and plants and in birds). Several studies have also found TBBPA in a variety of
environmental media that includes sediment, soil, landfill leachates, sewage sludge, surface water,
wastewater and indoor and outdoor air.
1A reactive use is one in which TBBPA is covalently bound with the substrate matrix. An additive use means that TBBPA is
not reacted with the substrate.
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Ecological toxicity studies include acute and chronic data for aquatic organisms (including sediment-
dwelling organisms), terrestrial plants, soil invertebrates, birds and amphibians. Epidemiological
studies and toxicity studies to inform human health hazard are also available. These include acute,
repeated-dose, reproductive, developmental, carcinogenicity, genotoxicity, irritation and sensitization
studies. Some information is also available on toxicokinetics.
EPA/OPPT followed Agency guidance during problem formulation and reviewed previous assessments,
such as those by the EU and Canada, to help inform the proposed assessment for TBBPA (and TBBPA-
bis(allyl ether)). Other published and unpublished data sources were also reviewed. During problem
formulation, EPA/OPPT identified available fate, exposure and hazard data, and characterized potential
exposures, receptors and effects.
To assist in targeting sources and exposure pathways of most concern, high-end exposure values from
existing risk assessments or more recent data were compared with toxicity values. For the
environmental scenarios, EPA/OPPT developed concentrations of concern (COCs) by applying
routinely-used uncertainty factors. For human health scenarios, EPA/OPPT calculated a health-
conservative toxicity value for developing uterine tumors using a recent carcinogenesis bioassay from
the National Toxicology Program.
Results of Problem Formulation
The results of this problem formulation are illustrated in the conceptual models and described by the
analysis plan that seeks to answer several assessment questions.
In summary, EPA/OPPT will conduct additional risk analysis of potential exposure to organisms
surrounding manufacturing facilities using concentrations of concern to determine risk quotients.
EPA/OPPT will also investigate potential exposure for workers and the general population/consumers
under the TSCA Existing Chemicals Program using existing data and methods.
Occupational risks will focus on workers within manufacturing and various processing facilities. For the
general population (with some exposure from consumer products), risks will be aggregated for
multiple oral pathways (direct oral ingestion or indirect ingestion from inhaled dust) and will focus on
individuals near manufacturing facilities and those farther away from such facilities. EPA/OPPT will
compare these exposure estimates to cancer and developmental effects benchmark values to evaluate
risk.
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1 INTRODUCTION
As a part of EPA's comprehensive approach to enhance the Agency's management of existing
chemicals, in March 2012 EPA identified a work plan of chemicals for further assessment under the
Toxic Substances Control Act (TSCA)2. After gathering input from stakeholders, EPA developed criteria
used for identifying chemicals for further assessment3. The criteria focused on chemicals that meet
one or more of the following factors: (1) potentially of concern to children's health (for example,
because of reproductive or developmental effects); (2) neurotoxic effects; (3) persistent,
bioaccumulative, and toxic (PBT); (3) probable or known carcinogens; (4) used in children's products; or
(5) detected in biomonitoring programs. Using this methodology, EPA/OPPT identified a TSCA Work
Plan of chemicals as candidates for risk assessment in the next several years. In the prioritization
process, tetrabromobisphenol A (TBBPA) was identified for assessment based on its use as a flame
retardant in epoxy resin circuit boards and in electronic enclosures in consumer uses, acute aquatic
toxicity and environmental persistence. Other chemicals were added to the cluster during this initial
prioritization based on considerations described in Section 1.1.
EPA/OPPT is performing risk assessments on chemicals in the work plan. If an assessment identifies
unacceptable risks to humans or the environment, EPA/OPPT will pursue risk reduction. The target
audience for this risk assessment is primarily EPA risk managers; however, it may also be of interest to
the broader risk assessment community as well as US stakeholders interested in TBBPA and related
chemicals. The information presented in the risk assessment may be of assistance to other federal,
state and local agencies as well as to members of the general public who are interested in the risks of
TBBPA and related chemicals.
The initial steps in EPA/OPPT's risk assessment development process, which is distinct from the initial
prioritization exercise, includes scoping and problem formulation. During these steps EPA/OPPT
reviews currently available data and information, including but not limited to, assessments conducted
by others (e.g., authorities in other countries), published or readily available reports and published
scientific literature. During scoping and problem formulation the more robust review of the factors
influencing initial prioritization may result in refinement - either addition/expansion or
removal/contraction - of specific hazard or exposure concerns previously identified in the prioritization
methodology.
This document includes the results of scoping and problem formulation for TBBPA and related
chemicals. In the initial prioritization and scoping stages EPA/OPPT determined which chemical(s)
would be included and what uses would be considered in the assessment. During problem formulation,
EPA/OPPT identified available exposure and hazard data, and characterized potential exposures,
receptors and effects. EPA/OPPT developed two conceptual models, Figure 2-1 and Figure 2-2, and an
analysis plan (Section 2.6.2) as a result of problem formulation.
2 http://www.epa.gov/oppt/existingchemicals/pubs/workplans.html
3 http://www.epa.gov/oppt/existingchemicals/pubs/wpmethods.pdf
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1.1 Scope of the Assessment
The four chemicals in the TBBPA and related chemicals assessment have the following general
structure:
Br
Figure 1-1: General Structure for TBBPA and Related Chemicals
Table 1-1 identifies the chemical names, Chemical Abstract Service Registry Numbers (CASRNs) and
definitions of R groups that are shown in the above structure.
Table 1-1: Chemical Names and Structures
Chemical Name
TBBPA
TBBPA-
bis(dibromopropyl ether)
TBBPA-bis(allyl ether)
TBBPA-bis (methyl ether)
CASRN
79-94-7
21850-
44-2
25327-
89-3
37853-
61-5
CAS Name
Phenol, 4,4'-(l-methylethylidene)bis[2,6-dibromo
Benzene, l,l'-(l-methylethylidene)bis[3,5-dibromo-
4-(2,3-dibromopropoxy)-
Benzene, l,l'-(l-methylethylidene)bis[3,5-dibromo-
4-(2-propen-l-yloxy)-
Benzene, l,l'-(l-methylethylidene)bis[3,5-dibromo-
4-methoxy-
R =
H
CH2(CHBr)2H
CH2CH=
CH2
-CH3
Of all brominated flame retardants, TBBPA has the highest US and global production volume (BSEF,
2014; EPA, 2014b). It is used as both an additive and reactive flame retardant4 and it is reacted to
produce more than 70 compounds (EPA, 2015).
In the initial prioritization process, EPA/OPPT considered structurally similar non-polymeric compounds
that could be assessed along with TBBPA. Two of the compounds, TBBPA-bis(dibromopropyl ether) and
TBBPA-bis(allyl ether), have additive flame retardant uses in addition to reactive uses. EPA/OPPT chose
these two compounds based on the assumption that additive uses lead to higher potential for
exposure than compounds with only reactive uses. EPA/OPPT included a fourth chemical, TBBPA-
bis(methyl ether), in the cluster based on studies that show it has routinely been found in the
environment. Its presence is likely to be a result of microbial transformation of TBBPA (George and
Haggblom, 2008).
4 A reactive use is one in which TBBPA is covalently bound with the substrate matrix. An additive use means that TBBPA is
not reacted with the substrate.
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1.2 Regulatory and Assessment History
The United States government, individual states and foreign governments have regulated TBBPA or
taken other actions with respect to TBBPA and some of the other cluster members. Primary risk
assessments are the European Union's environmental and human health assessments (EC, 2006, 2008)
and Canada's recent environmental and human health assessment (EC/HC, 2013).
1.2.1 Federal
At the federal level, companies report production volume and chemical use information for TBBPA,
TBBPA-bis(dibromopropyl ether) and TBBPA-bis(allyl ether) under the Chemical Data Reporting rule
(EPA, 2014b). TBBPA and TBBPA-bis(dibromopropyl ether) are produced in volumes greater than 1
million pounds per year and are thus considered high production volume (HPV) chemicals as defined by
EPA and industry prepared a test plan that was submitted to EPA under the HPV Chemicals Challenge
Program. TBBPA emissions are reported yearly to the Toxics Release Inventory (TRI) (EPA, 2012e).
1.2.2 State
The Oregon Department of Environmental Quality classifies TBBPA as a priority persistent pollutant as
part of its water quality program based on concerns related to persistence and chronic toxicity to fish
(Oregon_DEQ, 2010a, 2010c). Oregon also provides data on use, exposure pathways and releases for
TBBPA under this program (Oregon_DEQ, 2010b).
Washington lists TBBPA as a chemical of high concern based on human health effects5 and presence in
humans and accordingly provides some toxicity and exposure information under the Children's Safe
Products Act (WSDE, 2011, 2014b). Washington also lists information regarding the amount of TBBPA
contained in children's products, as reported to the Washington State Department of Ecology (WSDE,
2014a).
California lists TBBPA and TBBPA-bis(dibromopropyl ether) as priority chemicals for biomonitoring
identified by a scientific guidance panel (SGP) from a list of designated chemicals compiled using
Centers for Disease Control National Biomonitoring Program and other SGP recommendations
(California_Biomonitoring, 2014b, 2014c). No information is available regarding when biomonitoring
might begin However, California is not yet biomonitoring for these chemicals
(California_Biomonitoring, 2013, 2014a). In addition, TBBPA and TBBPA-bis(dibromopropyl ether) are
listed as candidate chemicals under the state's Safer Consumer Products regulations. Presence on this
list indicates that the chemical meets criteria for the initial priority products list specified in
regulations. TBBPA is listed based on bioaccumulation, endocrine toxicity, environmental persistence,
neurotoxicity, reproductive toxicity and other (undefined) toxicity. TBBPA would only be named as a
chemical of concern if it was part of a product-chemical combination that is listed as a priority product
(CalEPA, 2013).
5 However, it should be noted that a primary reference is for a formulated product called Saytex 111, not for TBBPA as a
monomer.
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Maine uses a tiered system to classify chemicals (Maine_DEP, 2013c). The first level includes roughly
1400 chemicals of concern including both TBBPA and TBBPA-bis(dibromopropyl ether) (Maine_DEP,
2013c). This list uses tools such as the Washington state list and EPA TRI information (Maine_DEP,
2013a). The second tier is comprised of 49 chemicals of high concern; TBBPA is the only chemical from
the TBBPA and Related Chemicals Cluster on this list (Maine_DEP, 2013b).
Minnesota lists TBBPA and TBBPA-bis(dibromopropyl ether) as chemicals of high concern based on
persistence, bioaccumulation and toxicity and references other authoritative lists (MDH, 2013a,
2013b). Among these authoritative sources, Minnesota refers to Washington state and EPA'sTRI
information.
States do not appear to regulate or provide information on either TBBPA-bis(allyl ether) or TBBPA
bis(methyl ether).
1.2.3 International
1.2.3.1 European Union
In the European Union (EU), TBBPA and TBBPA-bis(dibromopropyl ether) are registered chemicals
under the Registration, Evaluation, Authorisation and Restriction of Chemicals (REACH) regulation. This
means that companies (registrants) provided the European Chemicals Agency with information on the
uses, properties, hazards and potential risks for these substances (ECHA, 2015b). The EU has also
conducted human health and environmental risk assessments of TBBPA (EC, 2006, 2008).
The other two cluster members are pre-registered in the EU, which means companies will be providing
information on the properties and uses of these substances to the European Chemicals Agency by the
extended registration deadline of May 2018. These two substances are pre-registered because the
substances are manufactured or imported at 1 to 100 metric tons per year (ECHA, 2014).
On October, 2014, ECHA proposed to update the Community Rolling Action Plan (CoRAP) and identified
TBBPA as a proposed substance for evaluation in 2015. Member states evaluate substances included in
the CoRAP to determine if the measures in place are enough to manage the risks. If not, an evaluation
of whether there is a need for further action is conducted. Such action could include restrictions,
identification of substances of very high concern or other actions outside the scope of REACH. The final
decision regarding the substances to be evaluated in 2015 will be adopted by ECHA at the end of
March 2015 (ECHA, 2015a).
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1.2.3.1.1 European Union Risk Assessment
Environmental Assessment
In their environmental risk assessment (EC, 2008), the European Union concluded that there is a need
to limit risks to surface water, sediment and terrestrial organisms when TBBPA is used as an additive
flame retardant in formulation (compounding) of acrylonitrile butadiene styrene (ABS). Risks for the
terrestrial compartment were also identified at ABS conversion sites (which conduct final processing of
semi-finished products with TBBPA as an additive flame retardant); the risks were found only when
sewage sludge is applied to agricultural land. Risks to terrestrial organisms were also identified at sites
where TBBPA is reacted into epoxy/polycarbonate resins. In addition, the European Union concluded
there may be risks to organisms in the marine environment (EC, 2008).
The EU environmental assessment indicated TBBPA may degrade to bisphenol A (BPA) in anaerobic
sediments (EC, 2008). Formation of BPA from TBBPA was assessed in the updated risk assessment of
BPA; after testing in snails was conducted, no risks were reported for all environmental organisms
evaluated (United_Kingdom, 2008).
Human Health Assessment
The human health assessment of TBBPA (EC, 2006) identified no health hazards of potential concern to
adults. Therefore, no risk characterization was performed for workers. Also, because no health effects
were identified for adults and because consumer exposures were found to be negligible, the European
Union did not identify concerns for consumers. Similar conclusions were reached by the European
Union when evaluating risks to humans via the environment (through food, air and drinking water).
For infants, non-cancer health hazards were identified and two exposure scenarios were analyzed. One
exposure scenario was based on the environmental scenario used for adults (described above) and the
second scenario was based on exposure of infants via breast milk. The European Union concluded that
risks were low and that risk reduction measures were not needed (EC, 2006).
1.2.3.1 Canada
Pursuant to section 74 of the Canadian Environmental Protection Act, 1999, the Canadian government
identified TBBPA and TBBPA-bis(dibromopropyl ether) in the categorization of the Domestic
Substances List (DSL) as priorities for screening assessment because they met the criteria for
persistence and inherent toxicity to non-human organisms. TBBPA was determined to present an
intermediate potential for exposure of individuals in Canada.
On November 30, 2013, Canada published their final screening assessment of TBBPA and TBBPA-
bis(allyl ether) along with another substance, TBBPA bis(2-hydroxyethyl ether) (CASR RN 4162-45-2),
which is not included in the EPA cluster (EC/HC, 2013)Environment and Health Canada conducted both
environmental and human health assessments for TBBPA and TBBPA-bis(allyl ether) (EC/HC, 2013). For
environmental organisms, Canada concluded the quantity of TBBPA and TBBPA-bis(allyl ether) that
may be released to the environment is below the level expected to cause harm to organisms.
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For humans, conservative estimates of exposure suggested that breast-fed infants may be exposed to
more TBBPA than older Canadians. Yet, recent studies have shown that TBBPA was virtually
undetected in breast milk and blood samples from pregnant women in North America. Therefore,
Canada concluded that TBBPA is not harmful to human health at current levels of exposure. TBBPA-
bis(allyl ether) was also determined to not to be harmful for human health (EC/HC, 2013).
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2 PROBLEM FORMULATION
The problem formulation stage is intended to determine the major factors to be considered in the
assessment, including exposure pathways, receptors and health endpoints (EPA, 1998, 2014f).
Accordingly, this problem formulation document summarizes the exposure pathways, receptors and
health endpoints EPA/OPPT has recommended for inclusion in the risk assessment. To make this
determination, EPA/OPPT conducted a preliminary data review to identify available fate, exposure and
hazard data and determine its likely suitability for quantitative analysis. EPA/OPPT summarized the
outcome of this evaluation in conceptual models that illustrate the exposure pathways, receptor
populations and effects that will be considered in the risk assessment. EPA/OPPT also prepared an
analysis plan to demonstrate the proposed approach to address remaining defined assessment
questions that are possible based upon TSCA uses and best available data, tools and models.
TBBPA
Previous risk assessments identified several human health and some environmental scenarios that
resulted in low risks for TBBPA. Some environmental scenarios evaluated in these previous
assessments resulted in risks. Although EPA/OPPT referred to these previous risk assessments as much
as possible to inform the current problem formulation process, there are several updates that were
considered during the current effort that differ from the outcome of previous assessments.
Unlike Canada and the European Union, TBBPA is manufactured in the United States. Therefore
EPT/OPPT identified exposure near manufacturers for further evaluation. Also, a recent cancer
bioassay has been published (NTP, 2014a) and has been used to determine which human health
scenarios to evaluate further. Finally, new information regarding the presence of TBBPA in children's
products prompted EPA/OPPT to consider evaluating risks to children of certain ages as a result of
exposure to consumer products.
To assist in determining pathways that may be of particular concern, EPA/OPPT calculated very
preliminary estimates by comparing high-end exposure values from existing risk assessments or recent
published and unpublished data with provisional toxicity values. For the environmental scenarios,
EPA/OPPT developed preliminary concentrations of concern (COCs) by applying routinely-used
uncertainty factors (EPA, 2012d, 2013b) to results of selected ecotoxicity studies. For human health
scenarios, EPA/OPPT calculated a provisional health-conservative toxicity value for developing uterine
tumors using the recent carcinogenesis bioassay from the National Toxicology Program (NTP, 2014a).
Scenarios specific to young children were also compared with a no observed adverse effect level
(NOAEL) from a developmental toxicity study (Fukuda et al., 2004). These preliminary calculations
allowed EPA/OPPT to understand which pathways might be of highest concern for TBBPA.
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Other Cluster Members
EPA/OPPT concluded that no quantitative risk assessment is needed for cluster members other than
TBBPA for one or more of the following reasons: limited information, inability to use the more robust
data for TBBPA to read across to other cluster members, low toxicity or likely low risk concerns. More
information on these decisions and data for these chemicals are presented in Appendix A.
Appendix B presents information on literature searching and data adequacy, and Table 2-1 lists the
type of information needed for each scenario, population and type of hazard. The literature search will
be updated before conducting the risk assessment.
Table 2-1: Data Required for Risk Assessment
Workers
General
Population
Consumers
Ecological
Receptors
Exposure
Scenarios
Manufacture and
processing
Releases to the
environment
from
manufacturing
Consumer
product uses
resulting in
direct
exposures or
releases to
indoor
environments.
Releases to the
environment from
manufacturing
Exposure
Measured or modeled concentrations in relevant media may be used. A
combination of these approaches may be considered depending on the
receptor and exposure scenario of interest
Hazard/Toxicity
Hazard data, low dose extrapolation to obtain a
cancer slope, bioavailability
Acute and chronic
effects data
2.1 Physical and Chemical Properties
Table 2-2 identifies the physical-chemical properties for TBBPA, the only cluster member considered
for further evaluation. Estimation program values (EPA, 2013a) for vapor pressure and water solubility
were unrealistically low and log Kow values were unrealistically high. Therefore, estimated values are
not reported in Table 2-2.
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Table 2-2: Physical-Chemical Properties
Chemical Name
TBBPA
CASRN
79-94-7
Melting
Point (°C)
178
Boiling Point (°C)
> 200°C
(decomposes)
Vapor
Pressure
(mm Hg
@ 25 °C)
< 1 x 10-6
(a)
Water
Solubility
(mg/L)
4.16
@ 25°C(b'c)
Octanol-Water
Partition
Coefficient
(log Kow)
5.90
ND: no measured data
EC (2006)
(b'Morfetal. (2003)
(c) Highest measured value (at neutral pH) from available literature
2.2
Production Volume and Uses
This section describes production and use information that informs EPA/OPPT's discussions on sources,
pathways and receptors to be evaluated for TBBPA.
2.2.1.1 Production
The 2011 national production volume of TBBPA as reported to the CDR was approximately 120 million
pounds (EPA, 2014b)6. Production volume data for TBBPA is presented in Error! Reference source not
found.. Production volume by company is not available in the publically available CDR because either
no data were reported or companies filed confidential business information (CBI) claims (as noted in
Error! Reference source not found.). There is little or no public information on the production or
import volumes of final products treated with TBBPA-based flame retardants.
According to the Brominated Science and Environmental Forum (BSEF), TBBPA has the highest global
production volume of all brominated flame retardants (BSEF, 2014).
2.2.1.1.1
Manufacturers
The CDR data (EPA, 2014b) identifies five companies that manufacture and/or import TBBPA:
• Albemarle Corporation
• ICL Industrial Products (ICL-IP) America, Inc.
• LG Chemical America
• Sabic Innovative Plastics US, LLC
• A company that was claimed as CBI
6 The 2012 release of the CDR database contains domestic manufacturing volumes, import volumes, export volumes, and
industrial and consumer use data for reporting sites during the year 2011. In addition, it reports past production volume,
which includes domestic manufacturing volumes (including imports) for the year 2010.
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Table 2-3: CDR Production Volume Data
Company Site
Albemarle Corp South Plant
2270 Highway 79 South
Magnolia, AR 71753-9129
ICL-IP America, Inc.
622 Emerson Road, Suite 500
St. Louis, MO 63141-6742
LG Chemical America
910 Sylvan Avenue
Englewood Cliffs, NJ 07632
Sabic Innovative Plastics US, LLC
State Route 892
Washington, WV 26181-0068
CBI
Domestic
Manu-
facturing
CBI
ND
ND
ND
CBI
Imported
(Ibs)
ND
CBI
485,159
CBI
ND
Exported
(Ibs)
CBI
CBI
0
0
0
Used
on
Site
(lbs)a
CBI
N/A
N/A
N/A
0
2010 Past
Production
(Import/
Manufacture)
CBI
CBI
454,941
CBI
CBI
2011
National
Production
(Ibs/yr)
119,837,559
a The total volume (domestically manufactured and imported) of the chemical used at the reporting site. This number represents the
volume of the chemical that did not leave the manufacturing site.
CBI = Confidential business information
ND = No Data; the company did not provide the requested information.
N/A = Not Applicable; the imported chemical was never physically at the site
2.2.1.1.2 Trade Names
Albemarle Corporation and ICL-IP sell TBBPA as SAYTEX® CP-2000 and FR-1524, respectively
(Albemarle, 1999; ICL-IP, 2013b). Trade literature does not indicate the trade name of Sabic Innovative
Plastics US, LLC's or LG Chemical America's TBBPA products. Additionally, although not listed in the
2012 public CDR, company websites state that Chemtura7 sells TBBPA as BA-59P and Dover Chemical
Corporation sells the chemical as Doverguard 59 (Dover_Chemical, 2012; ICL-IP, 2013b). Chemtura
reported manufacturing TBBPA to the 2006 Inventory Update Rule (IUR), but Dover Chemical
Corporation did not (EPA, 2014h). For more detailed information on manufacturers of TBBPA who
reported for the CDR collection period that reported the 2011 production volume data, see Error!
Reference source not found..
2.2.1.1.3 Import and Export
CDR data indicate that TBBPA is imported but do not indicate whether TBBPA is exported (EPA, 2014b).
According to the CDR, LG Chemical America imported 485,159 pounds of TBBPA in 2011, while other
companies either did not report import and export volumes or claimed them as CBI.
7 The division of Chemtura that produces bromine-based products (including TBBPA) is called Great Lakes Solutions
(Chemtura, 2014).
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2.2.1.2 Uses
The CDR provides data on the industrial and consumer uses of TBBPA (EPA, 2014b). This data is
summarized in Table 2-4.
For the purposes of the CDR, "industrial use" means use at a site where one or more chemical
substances or mixtures are manufactured (including imported) or processed. "Consumer use" means
the use of a chemical substance or a mixture containing a chemical substance (including as part of an
article) when sold to or made available to consumers for their use.
2.2.1.2.1 Past Use as a Plasticizer
Additionally, TBBPA has been used as a plasticizer (NIEHS, 2002). However, this use has not been
documented in more recent years (EC, 2006; EC/HC, 2013; EPA, 2014b, 2014h).
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Table 2-4: Industrial and Consumer Use Data for TBBPA from the CDR
Company Site
Albemarle Corp South Plant
2270 Highway 79 South
Magnolia, AR 71753-9129
ICL-IP America, Inc.
622 Emerson Road, Suite
500
St. Louis, MO 63141-6742
LG Chemical America
910 Sylvan Avenue
Englewood Cliffs, NJ 07532
Sabic Innovative Plastics US,
LLC
State Route 892
Washington, WV 26181-
0068
CBI
Type of
Processing
Processing-
incorporation into
formulation,
mixture or reaction
product
Not Known or
Reasonably
Ascertainable
Processing-
incorporation into
formulation,
mixture or reaction
product
Processing-
incorporation into
formulation,
mixture or reaction
product
ND
Processing as a
reactant
Industrial Use Data
Sector
Plastics Material
and Resin
Manufacturing
Not Known or
Reasonably
Ascertainable
Computer and
Electronic Product
Manufacturing
Plastics Material
and Resin
Manufacturing
ND
All Other Basic
Organic Chemical
Manufacturing
Industrial Use
Flame
retardants
Not Known or
Reasonably
Ascertainable
Flame
retardants
Flame
retardants
ND
Flame
retardants
Percent of
Production
Volume
26
71
100
100
ND
100
Consumer Use Data
Consumer Use
Product
Category
Electrical and
Electronic
Products
Not Known or
Reasonably
Ascertainable
Electrical and
Electronic
Products
Plastic and
Rubber Products
not covered
elsewhere
ND
ND
Commercial or
Consumer Use
Commercial
Not Known or
Reasonably
Ascertainable
Both
Commercial
ND
ND
Percent of
Production
Volume
26
71
100
100
ND
ND
ND = No Data; the company did not provide the requested information.
Source: EPA (2014b)
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2.2.1.2.2 Flame Retardant and Other Uses
TBBPA is one of the most widely used brominated flame retardants and is used as both an additive and
reactive flame retardant (EPA, 2008a). Additive flame retardants are incorporated into polymers via
physical mixing and are not chemically bound to the polymer. Reactive flame retardants are
incorporated into polymers via chemical reactions at an early stage of manufacturing. Because
manufacturers can incorporate additive flame retardants into the product up until the final stages of
manufacturing, it is usually easier for them to use additive rather than reactive flame retardants.
Reactive flame retardants have a greater effect on the chemical and physical properties of the polymer
into which they are incorporated than do additive flame retardants (EPA, 2008a).
TBBPA has also been used as a chemical intermediate in the synthesis of other brominated flame
retardants (NIEHS, 2002).
As stated in Table 2-4, TBBPA's main consumer use categories as a flame retardant are 1) electrical and
electronic products and 2) plastic and rubber products not covered elsewhere. The category "plastic
and rubber products not covered elsewhere" means that products are not covered under any other
plastic or rubber product categories within the CDR. In electrical and electronic products, TBBPA is
primarily used in printed circuit boards in the following products: telecommunications equipment,
computers, industrial controls, remote controls, video recorders and electronics (EC, 2006; Qu et al.,
2013). TBBPA is incorporated into epoxy resins for printed circuit boards as a reactive flame retardant
at 15 to 17% by weight (EC, 2006).
Epoxy resins are also used to encapsulate electronic components (Morose, 2006). Electronic
component encapsulates, which incorporate TBBPA at 2% by weight, are used to protect products from
hazardous environmental conditions, such as moisture and dust. These products include plastic and
paper capacitors, microprocessors, bipolar power transistors and other components of electrical
equipment (EC, 2006; Morose, 2006).
With respect to TBBPA's use in plastics and rubber products, it is likely the majority of this use is in
electrical and electronic products. For example, a primary application of TBBPA is its use as an additive
flame retardant in acrylonitrile butadiene styrene (ABS) resins (a type of plastic). These ABS resins are
used in the enclosures or casings around electronics such as TV or computer monitor casings or
components in printers, fax machines, photocopiers, vacuum cleaners, coffee machines and
plugs/sockets. TBBPA is used in ABS and other plastics at 14 to 22% by weight, often in combination
with antimony trioxide (EC, 2006).
In addition to the CDR another dataset, which contains information on types of products that may use
TBBPA is Washington State's Children's Safe Product Act (CSPA) database8. Under CSPA,
manufacturers, importers or whole sale distributors of children's products sold in Washington are
required to report if their products contain a Chemical of High Concern to Children -one of which is
TBBPA (Washington_State_DEC, 2014). As of September 6, 2014, TBBPA has been reported for use as a
! https://fortress.wa.gov/ecy/cspareporting
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surface coating flame retardant in artists' accessories9. It has also been reported to be present as
synthetic polymer flame retardant in powered "viewing toys10", "toy/games variety packs11" and in
powered toy vehicles. Additionally, it is reported to be used as a flame retardant in textiles in baby
car/booster seats; baby carriers; baby play pens/dens and baby swings. The concentrations of TBBPA in
these products were reported as ranging from < 0.05 to > 1% (Washington State Department of
Ecology, 2014b). It is not clear from the database whether these products were manufactured in the
United States or imported as articles. The use of TBBPA in textiles is not reported in the publically
available CDR.
Non-flame retardant applications of TBBPA reported in Washington's database of children's products
include the chemical's use as an adhesive in jewelry craft supplies, as a pigment in powered non-ride
toy vehicles, as a stabilizer in clothing accessories and as a component of plastic resin or polymers in
toys. The concentrations of TBBPA in these products were reported as ranging from < 0.01 to > 1%
(WSDE, 2014a).
TBBPA has also been reported as a contaminant with no function in children's footwear, clothing,
personal accessories, arts and crafts, baby feeding products (i.e., baby bibs, according to a single
report) and bedding in the CSPA database. All contaminant concentrations were reported as less than
0.05%, with about half reported as less than 0.01% (WSDE, 2014a).
SAYTEX® CP-2000 (Albemarle's TBBPA product) is used as a reactive or additive flame retardant and is
usually used in combination with other additives, such as antimony trioxide (Albemarle, 1999; EC,
2006). Additionally, ICL-IP's FR-1524 is used in epoxy, polycarbonate and phenolic resins (ICL-IP,
2013b).
ICL-IP's Web site also states that TBBPA is "an important intermediate in the preparation of more
sophisticated flame retardants" (ICL-IP, 2013b). Chemtura and Dover Chemical Corporation state that
their TBBPA products, BA-59P™ and Doverguard 59, respectively, are intended for use in
"thermoplastic and thermoset resin systems." These resin systems include epoxy resins,
polycarbonates, ABS and high impact polystyrene (Chemtura, 2013; Dover_Chemical, 2012, 2013).
9 The product categories used by Washington State are defined by GS1 Global Product Classification Standards (WSDE,
2013). Artists' accessories are defined as any products that can be described/observed as an item designed to aid the
artistic painting process (GS1, 2015).
10 Includes any products that can be described/observed as a powered educational toy designed to entertain and encourage
learning by viewing changing scenes or patterns (GS1, 2015)
11 Includes any products that can be described/observed as two or more distinct Toys/Games products sold together which
exist within the schema but belong to different classes, that is two or more products contained within the same pack which
cross classes within the Toys/Games family.
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2.2.1.2.3 Regulatory Status and Future Trends Regarding Uses
No current state or federal legislation restricts the use of TBBPA (BSEF, 2013). The European Union's
REACH regulations require manufacturers and importers to report the properties and uses of
substances that they manufacture or import at or above one metric ton per year, including TBBPA
REACH does not, however, restrict the use of TBBPA or similar substances (ECHA, 2015b).
Although not a formal regulation, Underwriters Laboratories (UL) 94 is a predominant standard for the
level of flame retardancy within products that can influence the specific flame retardant or the amount
of a flame retardant that is used in electronic products. Compliance with the standard is not required
nationally. However, individual companies, trade associations or local governments may choose to
require electronic products under their jurisdiction to meet a UL 94 rating (ULJDES, n.d.).
2.2.1.3 Summary of Production and Use
TBBPA has the highest production volume of the brominated flame retardants. Five companies
reported a total of 120 million pounds manufactured or imported in 2011, according to CDR data. It is
used primarily as a reactive flame retardant, with uses as an additive flame retardant as well. Its use as
a flame retardant is primarily for electrical or electronic products, where it may be present at levels up
to 22%. A vast majority of printed circuit boards that meet stringent flame retardancy standards use
TBBPA to achieve these ratings. TBBPA may also be used as an intermediate in the production of other
flame retardants.
2.3 Fate and Transport
In air, TBBPA may undergo direct photolysis and indirect photo-oxidation, although low vapor pressure
limits the amount of TBBPA vapor in air. In water TBBPA does not hydrolyze but may undergo
photolysis. TBBPA is expected to have low volatility and low mobility in soil. Under anaerobic
conditions, TBBPA may biodegrade to BPA. It may also undergo microbial 0-methylation to form
TBBPA-bis(methyl ether). Based on available data, TBBPA is considered moderately to highly persistent
in water, soil and sediment and has a low bioaccumulation potential. Limited and uncertain data
preclude assessment of risks from TBBPA's degradation products as part of the proposed risk
assessment for TBBPA.
Appendix C includes details related to the fate and transport of TBBPA in air, water, soil and sediment
as well as information on degradation products.
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2.4 Exposures
2.4.1 Releases to the Environment
TBBPA is a solid, is manufactured in plants in Arkansas and is then incorporated into polymer matrices
either additively or reactively as noted in Section 2.2.1.2.2. Products in which TBBPA is used as a flame
retardant are largely electronics; TBBPA can be used reactively in printed circuit boards and also
incorporated into the acrylonitrile butadiene styrene (ABS) resins used in plastic housing of electronic
products. TBBPA may be released at various stages of the life cycle: manufacture, processing, disposal,
recycling and use. Toxics Release Inventory (TRI) data are available for manufacturers and processors.
Information on releases is described in Appendix D.
2.4.1.1 Chemical Manufacturing
Based on North American Industry Classification System (NAICS) codes reported by facilities to TRI
(EPA, 2012e) as well as production volume and use data reported to EPA under CDR (2011 reporting
year) (EPA, 2014b) and IUR (2005 reporting year) (EPA, 2014h), three facilities are likely to manufacture
all of the TBBPA that is produced domestically.
Of the years of available TRI reporting (2000 to 2012), two of these manufacturers were in operation
from 2000 to 2011. In 2012, one of these manufacturing sites did not report TBPPA releases, and a
different manufacturer reported emissions instead.
According to TRI reports (EPA, 2012e), manufacturers and processors reported stack emissions to air.
TBBPA is likely to be emitted as dust rather than vapor based on its low vapor pressure. Companies
also released TBBPA to landfills.
2.4.1.2 Processing
Several facilities process TBBPA in the United States. The following processing sectors have reported
TBBPA releases to TRI: chemicals, textiles, hazardous waste/solvent recovery, transportation
equipment, plastics and rubber, paper and computers and electronic products (EPA, 2012e). For the
years 2000-2012, processors emitted approximately 0.24 to 6.2% of the air releases reported by the
manufacturing site with the highest release for these years (EPA, 2012e).
2.4.1.3 Recycling
EPA's Office of Resource Conservation and Recovery (ORCR) estimated that approximately 25% of
electronic waste (e-waste) was collected for recycling in 2009 (EPA, 2013d). After collection, electronic
products that contain TBBPA might be reused, refurbished or recovered for their materials (EPA,
2011a). Plastics can be recovered by chemical, mechanical or thermal processes (Kang and Schoenung,
2005). Kang and Schoenung (2005) state that plastics recovered from electronics in the United States
are used primarily in plastic lumber, outdoor furniture and road materials. However, it is not clear
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whether this statement refers to plastic housing that contains flame retardants because the presence
of additives (such as flame retardants) in plastics can act as an obstacle to recycling (Kang and
Schoenung, 2005).
Generally, circuit boards with epoxy resins that contain TBBPA are shredded because they cannot be
re-melted to be used again. However, some boards are smelted as part of the recycling process to
recover precious metals (EC, 2006; Kang and Schoenung, 2005).
In the European Union, computer recycling companies often incinerate plastic housing from electronics
(EC, 2006). If US recyclers operate in a similar manner, plastics that contain TBBPA may be incinerated
with the potential for emissions.
Products collected for recycling can also be exported. Yet estimates of such exports can range widely.
Some estimates include 7% of the total monetary value of collected e-waste (ITC, 2013) to 50% of
amounts generated in the western United States (BAN/SVTC, 2002). Due to these and other
uncertainties, EPA/ORCR has not developed a method to estimate the total amount of e-waste that is
collected in the United States and subsequently managed and processed domestically or exported
(EPA, 2013d).
2.4.1.4 Disposal
2.4.1.4.1 Disposal from Manufacturing, Processing and Recycling Facilities
Manufacturers dispose of TBBPA in onsite and offsite landfills. The offsite landfills were primarily RCRA
Subtitle C hazardous waste landfills. One manufacturing facility also reported disposal of TBBPA in the
category of "other off-site landfills" for 2002. "Other landfills" are non-hazardous waste landfills that
may be regulated under a variety of other federal, state and local programs. Processors also reported
disposal to RCRA Subtitle C and other landfills (EPA, 2012e). Finally, computer recycling companies
usually landfill (or incinerate) plastic housing from electronics in the European Union (EC, 2006); these
practices might be applicable to the United States.
2.4.1.4.2 Disposal of Consumer Products Containing TBBPA
Products that contain TBBPA can be disposed in various ways. In 2009, electronic waste (TVs,
computers, peripherals, mice, keyboards and cell phones) totaled approximately 2.37 million short
tons, as estimated by EPA/ORCR. The amount of this waste that contains TBBPA is unknown. This
amount of electronic waste (also called e-waste) is about 1-2 percent of the total municipal waste
stream (EPA, 2013d). Approximately 75% of e-waste was disposed in landfills and by other methods
compared with 25% that was collected for recycling.
Electronic waste after use is typically sent to landfills (EPA, 2011a). Electronic waste can also be sent to
waste-to-energy incinerators (EPA, 2011a). Products that contain TBBPA can also be sent to municipal
incinerators (Borgnes and Rikheim, 2004). Furthermore, ash generated from incineration can also be
sent to landfills.
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General household waste and landfill leachates can be sent to WWTPs, which can discharge effluents
to water. Contaminants in the effluents can either remain in surface water or partition to sediments.
WWTPs also generate sewage sludge, which can be processed to yield nutrient-rich organic materials
called biosolids that can be applied to agricultural land as fertilizer (EPA, 2014J).
2.4.2 Presence in the Environment and Bio monitoring Data
Table 2-5 outlines the type of data available regarding presence in the indoor and outdoor
environment as well as biomonitoring data available for TBBPA. Other sections in this Chapter describe
results of individual monitoring studies likely to be most representative of individual exposure
pathways. Details regarding individual studies are described in detail in Supplemental Files 1
(Biomonitoring), 2 (Environmental and Wildlife Monitoring) and 3 (Residential Monitoring).
Table 2-5: Availability of Exposure Data for TBBPA
BIOMONITORING (HUMAN)
Blood
Breast Milk
Adipose Tissue
Placenta
Urine
•
•
•
HUMAN EXPOSURE
Dust ingestion
•
USGS NWIS DATA
Water
Suspended sediment
Solids
Biota
AIR
Ambient Air
Indoor Air
SOIL
INDOOR DUST
•
•
•
•
SEDIMENT
Freshwater
Marine
•
•
SLUDGE
amended soil
biosolids
landfill
sewage
•
•
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Table 2-5: Availability of Exposure Data for TBBPA
WATER
AIR AND
BIOTA
aquatic
drinking water
groundwater
leachate
precipitation
surface water
wastewater
WATER
deposition
avian
fish
animals (including shellfish)
terrestrial animals
vegetation
•
•
•
•
= some data available, US or international
2.4.3 Occupational Exposures
TBBPA is manufactured primarily at three plants in the United States. TBBPA is also processed at other
sites and electronic and other consumer products can be recycled. Thus, there is a potential for
exposure to workers in all of these sectors.
EPA/OPPT considers inhalation of dust to be the most important TBBPA exposure pathway for workers.
In particular, the inhalation of air-suspended dust (particulate matter) that is subsequently trapped in
mucous and moved from the respiratory system to the gastrointestinal tract (EPA, 2011b) may
contribute to exposures. This will be referenced in the current document as incidental ingestion of
inhaled dust. Dermal exposure is also possible but available data indicates that absorption is limited.
Information on concentrations of TBBPA and particles not otherwise regulated within workplaces
relevant to TBBPA is described in Appendix E.
2.4.4 General Population Exposures
The general population may be exposed to TBBPA due to its widespread detection in the indoor and
outdoor environment. TBBPA has also been detected in several human and fish biomonitoring studies.
The general population may be exposed to TBBPA through oral, inhalation or dermal exposure,
although aggregate oral exposure is the focus of this assessment. Data summaries and references are
available in Supplementary Files 1 through 3.
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2.4.5 Consumer Exposures
TBBPA is used as a flame retardant, primarily in electronic products (such as TVs and computers).
TBBPA has also been found in a variety of other products such as small plastic toys, jewelry and other
children's products as identified in recent studies (Di Napoli-Davis and Owens, 2013; Gallen et al.,
2014; Keller et al., 2014; van Bergen and Stone, 2014).
Direct contact with products may lead to exposures depending on the conditions of use, such as
frequency and duration of contact with the skin and subsequent hand to mouth or object to mouth
contact. These products may also contribute to variable levels within indoor dust and air depending on
the diversity of products present within a given building. A number of published studies have reported
levels of TBBPA in indoor air and dust (see summaries of these studies in Supplementary File z).
Note that although exposure from contact with consumer products is proposed for assessment in
Section 2.6, the types of exposures are primarily are not from contact of using the product directly
(e.g., using a keyboard for a computer). Thus, they are discussed in the context of the general
population (or of individuals living near manufacturing facilities).
2.5 Hazard Endpoints
2.5.1 Ecological Hazard
Ecotoxicity tests of aquatic and some terrestrial organisms exposed to TBBPA have evaluated a variety
of effects including survival, immobilization, growth rate/biomass/yield, reproduction, emergence,
growth and shell deposition. Depending on the type of organism, both acute and chronic studies are
available. Based on results from some of these studies, TBBPA can be considered to be hazardous to
the environment.
In aquatic studies, a range of values and effects have been identified. The most sensitive species and
effects from acute and chronic studies are reported here. TBBPA exposure by the marine diatom
Skeletonema costatum resulted in a 72-hr ECso of 0.09 mg/L, based on decreased growth (Walsh et al.,
1987). One of the lowest ecotoxicity endpoint values for a water-column species is the 96-hr ECso of
0.098 mg/L based on shell deposition in the Eastern oyster, a marine invertebrate species (SLS, 1989a).
In a 70-day study using the blue mussel, a MATC of 0.023 mg/L was calculated based on growth rate
and shell length (ACC, 2005). In fish, acute LCso values are all at or less than 1 mg/L, with the lowest
value reported as 0.4 mg/L in rainbow trout (GLCC, 1978b). A MATC of 0.22 mg/L was determined for
fathead minnows in a 35-day test (SLS, 1989c).
In sediment, several 28-day studies using worms, emergent flies or amphipods have been conducted,
with the lowest MATC reported as 117 mg/L, based on effects on reproduction of a freshwater
blackworm (Krueger, 2002a).
A range of terrestrial plants that include corn, cucumber, onion, ryegrass and tomato have been tested
in 21-day studies, with the lowest MATC of 32 mg/kg dry soil (ACC-BFRIP, 2002). One 21-day and two
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56-day studies have been conducted using terrestrial earthworms. The lowest MATC was 0.44 mg/kg
dry weight in soil in one of the 56-day studies (ACC-BFRIP, 2005a).
Avian studies are limited, and reproductive and endocrine effects were not observed in the adult quail
after in ovo exposure via injection of TBBPA into yolks (Berg et al., 2001; Halldin et al., 2001). In
tadpoles, endocrine-related effects were not seen up to 500 ug/L (Garber et al., 2001) but thyroid
hormone mediated gene expression was affected at 5.4 and 54 ug/L (Veldhoen et al., 2006).
See Appendix F for details related to these ecotoxicological studies.
2.5.2 Human Health Hazard
Available toxicokinetics data in rodents indicate that TBBPA is absorbed by the gastrointestinal tract,
metabolized and excreted in the feces with limited tissue retention, with a half-life in humans of 2
days. There is limited transfer of TBBPA to the fetus. The acute hazard concern is low via the oral,
dermal and inhalation routes. Also, many repeated-dose, reproductive and developmental toxicity
studies in rodents found no effects, some at doses > 1000 mg/kg-bw/day. TBBPA has tested negative in
genotoxicity studies. Yet, there is some concern for cancer of the uterus as well as hemangiosarcomas
and hemangiomas in all organs as observed in a cancer bioassay (NTP, 2014a). There is also a possible
concern for developmental effects at 200 mg/kg-bw/day based on slight kidney lesions in newborn rats
exposed to TBBPA. The lesions persisted after cessation of exposure, possibly due to immature
metabolic capability or kidneys of these rats (Fukuda et al., 2004). Another study found very slight
hepatocyte necrosis at 140.5 mg/kg-bw/day in offspring of female mice exposed to TBBPA during
gestation through weaning of the offspring (Tada et al., 2006).
Neurotoxicity and neurobehavioral effects have not been confirmed. One study found some potential
for hearing loss when dams and newborns were dosed (Lilienthal et al., 2008) but there are questions
about methods and uncertainty about which are the most relevant doses (e.g., both newborns and
dams were exposed to TBBPA). An acute study resulted in some neurobehavioral effects but didn't
show a dose-response (Nakajima et al., 2009). No consistent neurobehavioral changes were seen in
adolescents exposed to TBBPA (Kicinski et al., 2012).
Appendix G presents a more detailed discussion of human health endpoints considered for the
proposed assessment.
2.6 Results of Problem Formulation
2.6.1 Conceptual Models
During problem formulation, two conceptual models were developed to identify important sources,
pathways, receptors and effects. See Figure 2-1 and Figure 2-2 respectively, for the proposed
environmental and human health assessments. The scenarios that EPA/OPPT proposes to quantify for
TBBPA are identified using solid arrows. Dotted lines are used for scenarios that cannot be evaluated
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quantitatively due to lack of applicable or adequate data or information. Table 2-6 and Table 2-7
outline the scenarios that are being assessed and those that will not be assessed forTBBPA.
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Figure 2-1: Conceptual Model for the TBBPA Environmental Assessment
SOURCES
EXPOSURE PATHWAYS ECOLOGICAL RECEPTORS EFFECTS
1. Chemical
Manufacturing
Dust
3. Recycling
Sediment-
Dwelling
Organisms
4. Disposal:
• From Above
Sources
• Of Final Products
LEGEND
• Solid lines = Pathway can be quantified
• Dashed lines = Pathway not significant, uncertain, or not quantifiable
• Shaded boxes/ovals = Pathway proposed for risk assessment; exposure
and toxicity can be quantified
Toxicity
Acute Toxicity
Chronic Toxicity
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Table 2-6: Environmental Exposure Scenarios Considered for Assessment
#
1
2
3
Sources/Exposure Pathways
Chemical
manufacturing
Processing
Recycling
Air- Particulates
Soil
Water
Sediment
Air- Particulates
Soil
Water
Sediment
Air- Particulates
Soil
Water
Sediment
Ecological
Receptors
Birds
Soil-dwelling
organisms
Fish,
Invertebrates,
Algae
Sediment-
dwelling
organisms
Birds
Soil-dwelling
organisms
Fish,
Invertebrates,
Algae
Sediment-
dwelling
organisms
Birds
Soil-dwelling
organisms
Fish,
Invertebrates,
Algae
Sediment-
dwelling
organisms
Included in
Assessment?
NO
YES
YES
YES
NO
NO
NO
NO
NO
Rationale, Limitations and Uncertainties
Adequate toxicity data not available for avian species using inhalation
as the exposure route
Preliminary evaluation using data from 1977 for an Arkansas site and
preliminary concentrations of concern from key ecotoxicity studies
suggests there is a potential risk.
Although physical properties limit TBBPA concentrations in surface
water, TBBPA deposition from manufacturing facilities might be of
concern.
Preliminary evaluation using data from 1977 for an Arkansas site and
preliminary concentrations of concern from key ecotoxicity studies
suggests there is a potential risk
Air emissions from processing plants are much smaller than those from
manufacturing plants
Several factors result in significant uncertainty regarding use of existing
data from other countries to evaluate the risks from recycling:
• differences in environmental monitoring results among
countries
• Potential differences in recycling practices among countries
• unknown amount of e-waste exported from the United States;
and
• no current methods to assess the recycling process identified
by EPA/ORCR
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Table 2-6: Environmental Exposure Scenarios Considered for Assessment
#
4
Sources/Exposure Pathways
Disposal
Incinerators:
Air-particulates
WWTPs:
Biosolids and
water
Landfills
Ecological
Receptors
Multiple
Multiple
Included in
Assessment?
NO
NO
Rationale, Limitations and Uncertainties
Limited data are available and destruction of TBBPA is likely when
incinerated
Preliminary calculations using conservative assumptions regarding
application of biosolids to agricultural land and high-end exposure
values in water near a sewage treatment plant suggests low concerns
Hazardous waste landfills (Title C): The majority of TRI releases are to
hazardous waste landfills; controls are in place to limit exposure (e.g.,
sites are covered).
Other landfills: It is expected that only a small amount of TBBPA would
be available and mobility in soil is limited given physical-chemical
properties. (Note that some leachate analyses have shown TBBPA
concentrations associated with particulate matter.)
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Figure 2-2: Conceptual Model for the TBBPA Human Health Assessment
SOURCES
EXPOSURE PATHWAYS
HUMAN RECEPTORS
EFFECTS
1. Chemical
Manufacturing
2. Processing
[workers only]
3. Recycling
4. Disposal:
• From Above
Sources
• Of Consumer
Products
5. Use of Consumer
Products:
• Homes
• Offices
• Daycare Facilities,
Etc.
Surface
Water,
[Fish,
Sediment]
Consumers,
General
Population
LEGEND
• Solid lines = Pathway can be quantified
• Dashed lines = Pathway uncertain or not quantifiable
• Shaded boxes/ovals = Pathway proposed for risk assessment;
exposure and toxicity can be quantified
Cancer
Developmental
Toxicity
(newborns, young
children)
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Table 2-7: Human Health Exposure Scenarios Considered for Assessment3
#
1
2
3
Sources/Exposure Pathways'5
Chemical
manufacturing
Processing
Recycling
Dust
[unintended
exposure from
incidental ingestion
of inhaled
particles/dust]
Soil/plants/livestock
- dietary
Surface water -
fish ingestion
Dust
[unintended
ingestion]
Soil/plants/livestock
Surface water -
Fish ingestion
Dust
[unintended
ingestion]
Soil/plants/livestock
- dietary
Surface water -
fish ingestion
Human
Receptors
Workers
General
population
near facilities
Workers
General
population
near facilities
Workers
General
population
near facilities
Included in
Assessment?
YES
YES
NO
YES
YES
NO
NO
Rationale, Limitations and Uncertainties
Information on TBBPA and non-specific dust concentrations in air of
manufacturing sites suggests potential concerns.
TRI reports that manufacturers emit TBBPA to air; there is potential
concern for individuals who live near manufacturing facilities
Data on plant uptake from soil are limited and no data are available for
bioaccumulation into livestock
Although physical properties limit TBBPA concentrations in surface
water, TBBPA deposition from manufacturing facilities might be of
concern; EPA/OPPT will estimate bioaccumulation of TBBPA into fish
from the water column [but bioaccumulation from sediment-dwelling
organisms to fish is not available].
Information on TBBPA and non-specific dust concentrations in air of
processing sites suggests potential concerns.
Air emissions from processing plants are much smaller than those from
manufacturing plants. Therefore, these emissions will not be modeled.
Several factors result in significant uncertainty regarding use of existing
data from other countries to evaluate the risks from recycling:
• differences in environmental monitoring results among
countries
• Potential differences in recycling practices among countries
• unknown amount of e-waste exported from the United States;
and
• no current methods to assess the recycling process identified
by EPA/ORCR
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Table 2-7: Human Health Exposure Scenarios Considered for Assessment3
#
4
5
Sources/Exposure Pathways'5
Disposal
Consumer
Product Use
Incinerators:
Dust/[unintended
ingestion]
WWTPs:
Water/soil
Landfills:
Surface water -
fish ingestion
Products
Indoor dust
Human
Receptors
General
population
Children
General
population
[adults and
children]
Included in
Assessment?
NO
YES
YES
Rationale, Limitations and Uncertainties
Limited data are available and destruction of TBBPA is likely when
incinerated
Preliminary calculations using conservative assumptions regarding high-
end exposure values in water near a sewage treatment plant and
application of biosolidsto agricultural land and suggests low concerns
Hazardous waste landfills (Title C): The majority of TRI releases are to
hazardous waste landfills; controls are in place to limit exposure (e.g.,
sites are covered).
Other landfills: It is expected that only a small amount of TBBPA would
be available and mobility in soil is limited given physical-chemical
properties. (Note that some leachate analyses have shown TBBPA
concentrations associated with particulate matter.)
Data available on TBBPA concentrations and surface loadings in
products, including children's products, suggests some potential for
concern
Preliminary calculations suggest that risks from this pathway alone are
low, yet this pathway is included to assess aggregate exposure
aSome pathways in this table are not being formally assessed. However, when conducting aggregate risk assessments, there may be contribution from
one or more of these 'unassessed' pathways because EPA/OPPT will use data, such as TBBPA concentrations in outdoor dust levels or in fish eaten by the
general population, that do not have identified TBBPA sources; this table simply indicates the pathways that are not being assessed as major sources of
TBBPA exposure.
bExposure pathways that depend on either inhalation of vapor or dermal uptake are not included in this table or in the conceptual models. TBBPA has a
very low vapor pressure and therefore, exposure to vapor is negligible. Available information also suggests limited dermal uptake.
cDrinking water from different sources/pathways could contain TBBPA; however, data are not available for TBBPA in drinking water.
dFood other than fish is not assessed because it is the purview of other agencies.
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2.6.2 Analysis Plan
EPA/OPPT proposes assessing both environmental and human health risks for the scenarios outlined
below based on available data and modeling of exposure. In the final risk assessment, EPA/OPPT will
characterize the assumptions, limitations and uncertainties of the assessment that is generated.
Qualitative levels of confidence in the information used for the risk assessment will be discussed for
transparency.
2.6.2.1 Environmental Assessment near Manufacturing Facilities
EPA/OPPT proposes to conduct an environmental assessment that focuses on the following key
question:
1. Are air releases from two manufacturing facilities occurring at levels that would result in risk
to aquatic, sediment-dwelling or soil-dwelling organisms?
During problem formulation, EPA/OPPT identified high-end concentrations of TBBPA in environmental
media surrounding two US TBBPA manufacturing facilities (Pellizzari et al., 1978; Zweidinger, Cooper,
Erickson, et al., 1979; Zweidinger, Cooper, and Pellizzari, 1979). EPA/OPPT also considered other data
near brominated flame retardant manufacturers in China (Yang et al., 2012) when choosing exposure
scenarios of concern. Although the US data are from 1977, sediment concentrations ranged from
undetected to 330 mg TBBPA per kg sediment (Pellizzari et al., 1978; Zweidinger, Cooper, and
Pellizzari, 1979). In soil, concentrations ranged from undetected to 150 mg/kg in 1977 (Pellizzari et al.,
1978). One very high-end surface water concentration of 4.87 u.g/L was measured in a lake in China
(Yang et al., 2012), which is located near several brominated flame retardant producers.
From preliminary comparison of these high-end environmental concentrations with provisional COCs,
EPA/OPPT determined that there might be current risk concerns for ecological receptors surrounding
manufacturers of TBBPA given the likelihood that TBBPA persists in the environment.
EPA/OPPT plans to use estimated stack air release data (see Appendix D) for the two manufacturers
that reported the highest releases over 13 years of TRI reporting (EPA, 2012e) as inputs to the
AERMOD air deposition model (EPA, 20141). Using facility air releases and site-specific inputs (e.g.,
meteorology, terrain), EPA/OPPT can calculate high-end estimates of the yearly TBBPA depositions
onto water, sediment and soil for each year of TRI release data (from 2000 to 2012).
For the environmental risk assessment, the environmental concentrations estimated from TRI releases
can be compared with COCs to determine environmental risks for aquatic, sediment-dwelling and soil-
dwelling organisms, expressed as risk quotients (RQs). A review of available ecological toxicity data and
recommended COC values are presented in Appendix F.
EPA/OPPT may consider estimates from all years of TRI reporting and their associated environmental
concentrations to estimate risks, and could accomplish this in different ways. EPA/OPPT could add
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exposures over multiple years of TRI reports, could focus on high-end exposures for a given year or
could evaluate both. EPA/OPPT will assume no degradation of TBBPA in the proposed assessment.
2.6.2.2 Human Health Risk Assessment
EPA/OPPT proposes to conduct human health risk assessments that focuses on the following two key
questions:
1. Are there risks to workers from exposure to TBBPA through ingestion of suspended
particulates/dust during manufacturing and processing activities?
As noted previously, workers may be exposed primarily by ingesting TBBPA dust from the air at work
sites. During bagging or loading operations, dust may be generated within facilities that manufacture
TBBPA and workers may be exposed by ingesting particles that are inhaled from the air. Also, during
loading and unloading operations, dust may be generated within facilities that process or compound
TBBPA. Dust generated from unloading operations is expected to be pure TBBPA. However, dust
generated from compounding and loading operations is not expected to be pure TBBPPA. These two
types of dust (pure vs. mixture) may have differences in bioavailability.
EPA/OPPT will evaluate risks from occupational exposure at manufacturing and processing plants.
Methods to quantify incidental ingestion of inhaled dust that consider issues such as bioavailability
must be developed.
2. Are there risks from aggregate exposures for the general population based on ingestion of
suspended particles/dust from outdoor air; dust ingestion from indoor environments; fish
ingestion; and/or mouthing of objects containing TBBPA?
Based on past TBBPA concentrations in environmental media at manufacturing facilities and more
recent information on TBBPA emissions from manufacturers, there is potential concern for individuals
living near such facilities. EPA/OPPT will investigate exposure for these individuals. In addition,
EPA/OPPT will aggregate the relevant facility-specific exposures with other known exposures to TBBPA
that focus on the oral route, including incidental ingestion of particles/dust from air. Although some of
the additional exposures considered for aggregation are expected to be minor, EPA/OPPT is
nonetheless interested in how such aggregation will affect risk estimates. These exposure estimates
will be developed for adults and children as appropriate. Risks near facilities can then be compared
with risks estimates calculated for individuals living farther away from manufacturing facilities.
Suspended particles/dust in outdoor air. As noted in Section 2.6.2.1, EPA/OPPT plans to estimate
TBBPA air concentrations as a result of air releases at TBBPA manufacturing plants to the external
environment. EPA/OPPT will estimate risks as incidental ingestion of inhaled dust after emissions from
manufacturers, and EPA/OPPT will further consider methods to quantify such incidental ingestion.
There are also several studies in various countries that measured TBBPA concentrations in ambient air
in different types of environments (e.g., rural, urban) that may represent exposures away from
facilities (see Supplemental File 2 for studies of TBBPA concentrations in outdoor air).
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Dust from indoor environments. A wide range of studies have reported TBBPA in dust in a variety of
indoor environments that can be ingestion directly from the air or from hand-to-mouth transfer of
settled dust. EPA will review these studies and combine this information with age-specific activity
patterns and exposure factors to estimate this type of exposure.
Fish ingestion. EPA/OPPT also plans to estimate exposure from fish ingestion. As described in Section
2.6.2.1, air emissions deposited to water can be combined with estimates of TBBPA bioconcentration
into fish and fish ingestion rates for families of recreational anglers living near the manufacturing
facilities. There are also data available on concentrations of TBBPA in fish in various environments that
may be appropriate to assess risks for individuals who live farther away from manufacturing facilities
(see Supplemental File 2 for information on TBBPA concentrations in fish). More details on parameters
that can be used for evaluation of fish ingestion are located in Appendix H.
Mouthing of products by children. Young children are likely to exhibit higher exposure than older
children and adults due to their more prevalent object-to-mouth behavior. Therefore, EPA/OPPT will
assess ingestion of TBBPA by children from direct contact with objects and hands that have touched
such objects.
EPA/OPPT has found some data on both the concentrations and surface loadings of TBBPA in consumer
products to which routine contact is possible. This information can be combined with information from
Agency models and age-specific activity patterns and exposure factors to estimate exposure. More
details about the proposed assessment approach are described in Appendix I.
For both key questions, the exposure estimates will be compared against relevant toxicity benchmarks.
A cancer benchmark (uterine tumors for females; hemangiomas/hemangiosarcomas for males)
developed using a linear low-dose model and a developmental toxicity benchmark will be used as
appropriate for each exposure pathway (see Appendix J for the dose-response modeling of tumor data;
and Appendix G for hazard information).
2.6.3 Sources and Pathways Excluded from Further Assessment
Several sources and pathways were excluded from further assessment for lack of data or expected low
risks and are indicated in the conceptual models using dotted arrows.
2.6.3.1 Chemical Manufacturing
Environment. Exposure via directly inhaling TBBPA will not be assessed because no information is
available on the toxicity of tetrabromobisphenol A to plants and other wildlife organisms (e.g., birds)
exposed via the air.
Human Health. EPA/OPPT is not proposing to assess the potential for dietary intake from eating crops
and livestock around manufacturers for several reasons. Although the EU Risk Assessment (EC, 2008)
used Kow instead of measured bioconcentration or bioaccumulation factors and Koc values were used
to determine uptake to plants and then livestock, more recent data of the uptake of TBBPA by cabbage
and radishes from soil showed that a large amount of TBBPA was adsorbed to soil and not available for
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transfer to plants (Li et al., 2011). Second, no data were found regarding the bioaccumulation of TBBPA
into livestock. Third, the evaluation of exposures from food other than fish is the purview of agencies
other than EPA. Appendix K presents details of the EU scenarios that resulted in risks to the
environment.
2.6.3.2 Processing
The EU risk assessment for the environment found risks surrounding both acrylonitrile butadiene
styrene (ABS) compounding sites and ABS conversion/epoxy resin manufacturing facilities from
disposal to waste water and subsequent application to agricultural land or from air releases (EC, 2008).
However, in the category of US plastics and rubber processors reporting to TRI, very minimal releases
to wastewater and air were reported, with none reported as being used for treatment on agricultural
land (EPA, 2012e).
US processing sites in sectors other than the plastics and rubber sector have reported higher stack air
releases to TRI (EPA, 2012e) than for the plastics/rubber sector. However, EPA/OPPT does not propose
evaluating these releases either because they are only a small proportion of the air emissions from
manufacturing sites.
Appendix K presents details of the EU scenarios that resulted in risks to the environment.
2.6.3.3 Recycling
Environment. TBBPA has been found in areas surrounding electronics recycling facilities in China, in
soil, sediment and nearby waterways (Xu et al., 2012; Feng et al., 2012; He et al., 2010). In contrast,
Schlabach et al. (2011) did not find TBBPA in sludge near car demolishing, waste recycling and
municipal recycling/landfills in Norway. See Supplemental File 2 for details on TBBPA concentrations
near recycling facilities.
There is significant uncertainty in evaluating the risks from recycling: consumption of TBBPA is
expected to be higher in Asia than in the United States (He et al., 2010) whereas TBBPA was not
detected in sediment and sludge in Norway; regulations for handling e-waste may differ between
countries; a significant (yet unknown) amount of e-waste may be exported from the United States; and
methods to assess other aspects of the recycling process have not been identified by EPA/ORCR.
Because of these uncertainties, EPA/OPPT will not evaluate risks from TBBPA present in environmental
media surrounding recycling facilities.
Human Health. Workers at recycling plants may be exposed to TBBPA particulates. Based on an
assessment of computer and plastic recycling operations, the European Union estimated typical and
reasonable worst case exposures as 0.02 and 4 mg/m3, respectively. The highest exposure potential
was associated with plastic recycling (EC, 2006). Also, TBBPA concentrations were found in
environmental media near e-waste recyclers as noted in the previous section, and these
concentrations could affect the general population living near such facilities.
EPA/OPPT is not planning to evaluate risks for workers or the general population given significant
uncertainties regarding the recycling process in the United States as defined by EPA/ORCR.
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2.6.3.4 Disposal
2.6.3.4.1 Landfills
Some waste that contains TBBPA may be classified as hazardous and must be sent to RCRA Subtitle C
hazardous waste landfills. EPA regulations at such landfills would significantly limit any exposure from
these off-site disposals. Controls include double liners, double leachate collection and removal
systems, a leak detection system, along with additional measures (EPA, 2013c). When non-hazardous
waste is disposed in properly-managed municipal solid waste landfills (which also must comply with
regulations to limit exposure), e-waste is not expected to threaten human health or the environment
according to EPA/ORCR (EPA, 2013d).
Furthermore, only limited leaching of TBBPA from landfills is likely because TBBPA is expected to
adsorb to soil particles (based on its log Koc of 5.4). In their assessment, Canada noted the limited
potential for TBBPA to reach groundwater (EC/HC, 2013). TBBPA has been measured in leachates from
landfills in the Netherlands, Finland and Japan (de Boer et al., 2002; Osako et al., 2004; Peltola, 2002;
Suzuki and Hasegawa, 2006). Most often TBBPA concentrations are quite low. Prior to treatment,
however, TBBPA may be found at higher concentrations (up to 320 u.g/kg dry weight), as seen in the
Netherlands (de Boer et al., 2002); data on pre- and post-treatment suggests that concentrations could
decrease by > 88 to 98% after treatment (Osako et al., 2004). Supplemental File 2 presents data on
measured TBBPA levels in leachates and other environmental monitoring studies.
For the above reasons, EPA/OPPT will not evaluate risks from disposal of final products after use for
the environment or humans.
Landfills that are no longer in operation or that are out of compliance with regulations limiting releases
may result in the potential for exposure. However, an evaluation of these situations is beyond the
scope of the proposed assessment.
2.6.3.4.2 Incinerators
EPA/OPPT found only one study measured TBBPA emissions (0.008 ng/L to air) from a mixed
household and commercial waste incinerator in Japan (Borgnes and Rikheim, 2004). Also, EC/HC
(2013) assumed that control devices on incinerators would limit releases of TBBPA to air. Therefore,
due to limited data and likely destruction of TBBPA during incineration, EPA/OPPT will not calculate
risks from incineration of TBBPA-containing products for the environment or humans.
2.6.3.4.3 Wastewater Treatment Plants (WWTPs}
Facility waste and final consumer products that contain TBBPA may be sent to WWTPs. Exposure to
TBBPA could occur after discharge of effluents from WWTPs to water, where it could remain in surface
water or partition to sediments or from generation of sludge that is then applied to agricultural land.
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Environment- Water. TBBPA has been found in WWTP effluents, waterbodies, sediment and sewage
sludge/biosolids (see Supplemental File 2 for individual study data). Comparing a high-end surface
value of 0.02 u.g/L downstream from a sewage treatment plant in Germany (Kuch et al., 2001) with a
chronic COC of 1 u.g/L suggests low concern. Also, previous assessments have not identified risks from
disposal of consumer products (EC, 2008; EC/HC, 2013).
Environment - Sludge Applied to Land. EPA/OPPT conducted a preliminary calculation to estimate
potential TBBPA concentrations in soil after application of sewage sludge to land using conservative
assumptions (no degradation and ten years of application) and a high-end TBBPA concentration in
sludge (1329 u.g/kg) from Spain (Guerra et al., 2010). The resulting TBBPA concentration after soil
mixing is below the COC determined for earthworms (= 44 u.g/kg), suggesting low concern. North
American levels in sludge (Quade, 2003), although older by a few years, are all lower than Guerra et al.
(2010).
Human Health -Water. It is possible that individuals may eat fish or obtain drinking water in areas near
WWTPs. Data on TBBPA concentrations in fish were not located for areas specifically located near
WWTPs.12 Also, data on TBBPA concentrations in treated drinking water is not available. Therefore,
EPA/OPPT will not assess these pathways in the current assessment.
However, as a preliminary exercise to determine whether there are possible concerns for people in the
vicinity of WWPTs, EPA/OPPT used an exposure estimate from the Canadian assessment (EC/HC, 2013)
based on a high-end surface water TBBPA concentration of 0.02 u.g/L from Kuch et al. (2001) obtained
near a WWTP; the highest exposure estimate from EC/HC (2013) was 1.0 x 10~6 mg/kg-bw/day, for ages
0.5 to 11 years. EPA/OPPT multiplied this exposure value by the most health-conservative slope factor
from the NTP (2014a) bioassay.13 The preliminary risk of 3.3 x 10~9 is approximately 300 times lower
than the target risk level of 1 x 10'6.14
Note that other surface water values in non-manufacturing areas were lower than the values in Kuch
et al. (2001). Therefore, based on information from current published studies on TBBPA in surface
waters, risk from TBBPA in drinking water is likely to be of low concern for non-industrial areas.
Although sewage sludge can be applied to agricultural land, risks resulting from this possible scenario
are not being considered for lack of information on uptake from soil, as described in Section 2.6.3.1.
2.6.3.5 Other Excluded Pathways
12 Fish ingestion for the general population not specific to this exposure pathway will be included in the risk assessment as a
part of the aggregate risk assessment.
13 Value is 0.00329/[mg/kg-bw/day] for development of uterine tumors
14 The age group of 0.5 to 11 years combines two ranges (0.5-4 years; 5-11 years) from EC/HC (2013) and was used because
several years of exposure are most appropriate (e.g., roughly l/10th of the human life span) when comparing with the
cancer benchmark. The resulting estimated risk is calculated as 0.00329/[mg-kw/bw/day] * 0.000052 mg/kg-bw/day = 1.7 x
10"7 and is lower than the target risk level of 1.0 x 10"6.
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Mouthing Products with TBBPA for Adults. Although children may place products that contain TBBBPA
in their mouths frequently, this is not considered a pathway of concern for adults, given the low
likelihood of such behavior for this population.
Drinking Water. As noted in Section 2.6.3.4.3, intake of TBBPA from drinking water is not being
assessed as part of the aggregate exposure estimates because data on TBBPA concentrations in treated
drinking water are not available. Yet, a preliminary calculation of specific to WWTPs suggests this
pathway may be of low concern.
Food. Because the presence of TBBPA in food is the purview of other agencies, it will not be evaluated
in the proposed TBBPA risk assessment, except that the ingestion of fish by the general population will
be used as a comparison to fish intake for individuals who live near manufacturing facilities.
2.6.4 Uncertainties and Data Gaps
2.6.4.1 Environmental Fate Data
Biodegradation data are lacking for organisms that have not already been acclimated to TBBPA. In
addition, for biodegradation and photolysis endpoints, data on rates of transformation and identity of
degradation products are limited.
2.6.4.2 Release Data
The TRI database is a comprehensive source of environmental release data for the United States.
However, there are certain limitations and uncertainties when using the data for a risk assessment.
For example, TRI information is self-reported and limited to those facilities that meet certain criteria
(EPA, 2012e). A facility must report to TRI if it:
1. is in a specific industry;
2. employs 10 or more full-time equivalent employees; and
3. manufactures or processes more than 25,000 pounds of a TRI-listed chemical or
otherwise uses more than 10,000 pounds of a TRI-listed chemical in a given year.
In addition, facilities can use various methods to estimate the releases they report to TRI (EPA, 2012e).
These methods can include continuous monitoring, periodic monitoring, and use of emission factors,
best engineering judgment and other methods.
Over the past 13 years, the two facilities considered for quantitative risk assessment - Great Lakes
Chemical Solutions (with Chemtura as the parent company) and Albemarle - have used engineering
calculations, published emission factors or site specific emission factors as a basis for reporting their air
emissions to TRI.
2.6.4.3 Exposure Information
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Several exposure parameters to be included in any modeling for TBBPA are variable and limited data
exist. Therefore, there is uncertainty in evaluating risks for children exposed through contact with
consumer products and subsequent ingestion of TBBPA. Furthermore, there is a lack of current
measured TBBPA concentrations in environmental media. There is a lack of measured TBBPA
concentrations in sediment in a variety of areas in the United States. Finally, very limited data are
available on TBBPA concentrations in the workplaces specific to manufacture and use of TBBPA.
2.6 A A Ecological Hazard Data
Uncertainties and limitations of the hazard data may result from the lack of robust data for aquatic,
sediment, terrestrial and avian species. Most importantly, exposure to TBBPA will likely occur primarily
in sediments and soil. Yet, few acceptable sediment and soil toxicity experiments have been conducted
on TBBPA.
Uncertainties also exist in the assessment factors typically used with toxicity values to determine
concentrations of concern. Actual variability may differ from the values of 4 or 5 for acute studies and
10 for chronic studies typically used by EPA/OPPT for TSCA-related activities.
2.6.4.5 Human Health Hazard Data
NTP concluded that for hepatoblastomas in male mice, there is some evidence of carcinogenic activity
that can be attributed to TBBPA. However, the data did not provide a good fit using the cancer
multistage model (Hummel, 2013b), and none of the other models available in the benchmark dose
response modeling software resulted in goodness of fit p-values at acceptable levels of > 0.1 when
considering the full shape of the dose-response curve (Hummel, 2013a). Thus, this tumor type cannot
be considered in a quantitative risk assessment of TBBPA.
The CARC noted that according to the IPCS MOA framework (IPCS, 2007), data are not adequate to
draw conclusions about the mode(s) of action for the tumor incidence associated with TBBPA.
Therefore, as recommended by EPA (2005), linear low dose extrapolation was used as the default
option for modeling tumor data. Yet, because mode of action data are not conclusive, it is possible that
a non-linear mode of action could explain the relationship between TBBPA and tumor incidence.
Cancer multistage models were chosen for modeling the cancer bioassay data based on adequate fits
for two tumor types and biological considerations even though other models also resulted in adequate
fits of the data.
NTP (2014a) administered TBBPA to rodents via oral gavage throughout their lifetime at doses from
250 to 1000 mg/kg-bw/day. Thus, there is uncertainty as to whether less than lifetime or even lifetime
exposure by humans exposed to lower doses of TBBPA associated with dust particles or as other forms
for less than a lifetime or even a full lifetime. Data are not available to determine quantitative TBBPA
disposition for time periods longer than those evaluated in available toxicokinetics studies.
The studies evaluating reproductive and developmental toxicity show a wide variety of results from no
effects up to very high doses to subclinical effects at low doses when using TBBPA as the test
Page 47 of 135
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substance. Thus, there is uncertainty in choosing any developmental toxicity study for evaluation in a
quantitative risk assessment of TBBPA.
Only few inhalation and dermal studies are available and therefore, there is uncertainty as to effects
specifically from these routes. There are also only limited studies in humans.
Page 48 of 135
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Appendix A Data Available for TBBPA-bis(dibromopropyl ether),
TBBPA-bis(allyl ether) and TBBPA-bis(methyl ether)
Although EPA decided not to conduct assessments for the three cluster members other than TBBPA,
there are some data available for these compounds. Table_Apx A-l presents an overview of the data
available for all cluster members, to show the comparison of the other cluster members with the index
chemical TBBPA. The ability to use TBBPA data to characterize these other chemicals (i.e. the ability to
read-across) is also discussed for environmental and human health endpoints.
Table_Apx A-l: Data Availability and Read Across for Cluster Members
Endpoint
Chemical
TBBPA
TBBPA-bis(dibromo
propyl ether)
TBBPA-
bis(allyl ether)
TBBPA-
bis(methyl ether)
Production Volume and Uses
Production Volume
(2010 or 2011)
Industrial Uses
Consumer Uses
X
X
X
X
X
X
X
ND
ND
ND
ND
ND
Physical-Chemical Properties
Melting Point
Boiling Point
Vapor Pressure
Water Solubility
Octanol-Water
Partition Coefficient
X
X
X
X
X
X
X
X
ND
ND
X
X
X
ND
ND
X
X
X
ND
X
Fate Properties
Biodegradation
Experimental
Bioconcentration
Data
X
X
X
ND
ND
ND
ND
ND
Exposure (Monitoring Data)
Environmental
media
Biota
X
X
X
X
X
X
X
X
Environmental Effects
Acute fish
Chronic fish
Acute invertebrate
Chronic
invertebrates
Algae
Sediment toxicity
Earthworms
X
X
X
X
X
X
X
X
ND
X
ND
X
ND
X
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
Page 71 of 135
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Table_Apx A-l: Data Availability and Read Across for Cluster Members
Endpoint
Potential for read
across from TBBPA?
Chemical
TBBPA
NA
TBBPA-bis(dibromo
propyl ether)
No;
Physical-chemical
properties differ
TBBPA-
bis(allyl ether)
No;
Physical-chemical
properties differ
TBBPA-
bis(methyl ether)
No;
Physical-chemical
properties differ
Human Health
Acute Oral
Acute Dermal
Acute Inhalation
Repeated-Dose
Developmental
Reproductive
Genetic Toxicity
Neurotoxicity/
neurobehavioral
Skin Irritation
Eye Irritation
Sensitization
Carcinogen icity
Potential for read
across from TBBPA?
X
X
X
X
[oral and limited
dermal/inhalation]
X
X
X
[gene mutations,
chromosomal
aberrations, other]
X
X
X
X
X
NA
X
X
X
X
[oral]
ND
ND
X
[gene mutations,
chromosomal
aberrations, other]
ND*
X
X
X
ND
No;
This compound has
alkylating potential
[high mw, low
solubility limit
toxicity]
X
X
ND
ND
ND
ND
X
[gene mutations]
ND
ND
ND
ND
ND
No;
This compound has
alkylating/epoxide
forming potential
[high mw, low
solubility limit toxicity]
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
Yes;
Structurally and
mechanistically
similar to TBBPA
NA = not applicable; X = data available; ND = no data available
*in vitro only
The data for each of these endpoints are described in the following sections:
A-l. Exclusion from Further Assessment
A-2. Chemical Structures
A-3. Physical-Chemical Properties
A-4. Production Volumes
A-5. Uses
A-6. Fate Properties
A-7. Exposure
A-8. Ecological Hazard
A-9. Human Health Hazard
Page 72 of 135
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A-l Exclusion from Further Assessment
EPA/OPPT does not propose to conduct risk assessments for the three cluster members other than
TBBPA for reasons detailed below.
A-l-1 TBBPA-bis(dibromopropyl ether)
Regarding ecological toxicity, the only adequate data for this compound are for earthworms, and the
results indicate low toxicity (ECHA, 2013). The toxicity estimation program EcoSAR v. 1.11 (EPA, 2012b)
is not well suited for use with this compound to fill additional endpoints because log Kow values are
expected to be higher than those used to develop the algorithms used within EcoSAR. Yet, TBBPA-
bis(dibromopropyl ether) is likely to have low water solubility and a high octanol-water partition
coefficient and no toxic effects are expected in aquatic and other organisms when considering
quantitative structure activity relationships (QSARs) (Lipnick, 1995).
For human health, a toxicokinetics study in rats shows low absorption by TBBPA-bis(dibromopropyl
ether) (Knudsen et al., 2007), which limits its ability to reach target cells. Lack of absorption is expected
given the compound's high molecular weight and log Kow and low water solubility. Low toxicity has
been indicated in available studies and is likely to be due to its limited absorption (ECHA, 2013; GLCC,
1982; IPCS, 1995).
Due to differences in physical-chemical properties compared with TBBPA and its potential to act as an
alkylating agent, EPA/OPPT has determined that TBBPA toxicity studies should not be used as
surrogate data for this compound for either ecological or human health toxicity endpoints. Therefore,
EPA/OPPT will not assess risks to this compound for these reasons as well as the limited toxicity data
available. The expected low toxicity of the compound also suggests that gathering additional data on
this compound is not a high priority.
A-l-2 TBBPA-bis(allyl ether)
No adequate ecological toxicity data were found for TBBPA-bis(allyl ether) and it is not expected to
exhibit effects in aquatic and other organisms due to physical-chemical properties and QSARs (Lipnick,
1995).
Regarding the potential to result in human health effects, absorption is likely to be limited as is the
ability to reach target cells. Limited toxicity information suggests low toxicity (Abbott et al., 1981;
Brusick, 1977; EC/HC, 2013; Qu et al., 2011).
Similar to the dibromoether, differences in physical-chemical properties and reactivity due to TBBPA-
bis(allyl ether)'s alkylating potential preclude using TBBPA toxicity data to read across to this
compound. EPA/OPPT does not propose to assess risks from this compound due to these differences
and the limited toxicity information. Again, the expected low toxicity indicates that this compound is
not a high priority for obtaining data in order to conduct a risk assessment.
Page 73 of 135
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A-l-3 TBBPA-bis(methyl ether)
No adequate ecological toxicity data were found for this compound, and like the other two cluster
members, it is not expected to exhibit effects in aquatic and other organisms due to physical-chemical
properties and QSARs (Lipnick, 1995). Also, differences in physical-chemical properties compared with
TBBPA suggest that TBBPA cannot be used as a surrogate for ecological toxicity.
No human health toxicity data were found for TBBPA-bis(methyl ether). Unlike the ecological toxicity
endpoints, TBBPA could be used as a surrogate compound for TBBPA-bis(methyl ether) to estimate
human health hazards based on enough similarity in the chemicals' structures. However, TBBPA-
bis(methyl ether) is found only as a transformation byproduct of TBBPA; EC (2008) indicates that it may
be a minor degradation product. In addition, its concentrations in the environment are usually lower
than TBBPA concentrations (see Supplemental File 2 for environmental concentrations). Furthermore,
scenarios of concern for TBBPA are not relevant for TBBPA-bis(methyl ether) because it is not
manufactured or found in consumer products. Therefore, EPA/OPPT does not propose conducting a
risk assessment on this compound.
A-2 Chemical Structures
Individual compound structures for the three cluster members are given in the sections below.
A-2-1 TBBPA-bis(dibromopropyl ether)
A-2-2 TBBPA-bis(allyl ether)
Page 74 of 135
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A-2-3 TBBPA-bis(methyl ether)
A-3 Physical-Chemical Properties
TBBPA-bis(methyl ether), TBBPA-bis(allyl ether) and TBBPA-bis(dibromopropyl ether) are solids with
low vapor pressure and low water solubility. Table_Apx A-2 identifies the physical-chemical properties,
which are only available for some cluster members. Estimation program values (EPA, 2013a) for vapor
pressure and water solubility were unrealistically low and log Kow values were unrealistically high.
Therefore, estimated values are not reported in Table_Apx A-2. However, compared with TBBPA, the
other cluster members are expected to have lower water solubility and vapor pressures and higher
octanol-water partition coefficients (log Kows) given their larger size and higher molecular weights.
Page 75 of 135
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Table_Apx A-2: Physical-Chemical Properties
Chemical Name
TBBPA-
bis(dibromopropyl
ether)
TBBPA-bis(allyl
ether)
TBBPA-bis (methyl
ether)
CASRN
21850-
44-2
25327-
89-3
37853-
61-5
Melting
Point (°C)
95
120
ND
Boiling Point (°C)
> 200°C (dec)
> 200°C (dec)
>200°C(dec)
Vapor
Pressure
(mm Hg
@ 25 °C)
< Ix ID'6
< 1 x ID'6
< 1 x ID'6
Water
Solubility
(mg/L)
ND
ND
ND
Octanol-Water
Partition
Coefficient
(log Kow)
ND
ND
> 6.4
ND: no measured data
Source: IPCS (1995)
A-4 Production Volumes
A-4-1 TBBPA-bis(dibromopropyl ether)
According to the 2012 CDR, the production volume of TBBPA-bis(dibromopropyl ether) is between 1
and 10 million pounds per year, as shown in
Table_Apx A-3. The CDR database identifies ICL-IP America, Inc., ICC Chemical Corporation and one
company that claimed CBI as manufacturers/importers of TBBPA-bis(dibromopropyl ether) (EPA,
2014b).
Trade names for TBBPA-bis(dibromopropyl ether) include ICL-IP's FR-720 and Dover Chemical
Corporation's (a subsidiary of ICC Industries, Inc.) Doverguard 68 (Dover_Chemical, 2013; ICL-IP,
2013a).
CDR data do not indicate whether this chemical is imported (EPA, 2014b). All companies listed in the
2012 CDR database exported 0 pounds and did not report import volumes or claimed the data as CBI.
A-4-2 TBBPA-bis(allyl ether)
CDR data identifies ICC Chemical Corporation as a manufacturer/importer of TBBPA-bis(allyl ether).
The national production volume for TBBPA bis(allyl ether) is withheld from the 2012 CDR to protect CBI
claims. However, one of the two ICC sites reported a past production volume (for the year 2010) of
124,575 pounds (EPA, 2014b), as shown in
Table_ApxA-3.
Page 76 of 135
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CDR data do not indicate whether this chemical is imported (EPA, 2014b). All companies listed in the
2012 CDR database exported 0 pounds and did not report import volumes or claimed the data as CBI.
A-4-3 TBBPA-bis(methyl ether)
There are no reports for TBBPA-bis(methyl ether) in the 2012 CDR database (EPA, 2014b).
Table_Apx A-3: 2012 CDR Production Volume Data (Pounds/Year) for TBBPA-bis(dibromopropyl
ether), TBBPA-bis(allyl ether) and TBBPA-bis(methyl ether)
Chemical
Name
TBBPA-
bisfdibrom
o-propyl
ether)
TBBPA-
bisfallyl
ether)
TBBPA-
bis( methyl
ether)
Company Site
ICL-IP America, Inc.
622 Emerson Road,
Suite 500
St. Louis, MO 63141
ICC Chemical
Corporation
460 Park Avenue
New York, NY 10022
CBI
ICC Chemical
Corporation
460 Park Avenue
New York, NY 10022
ICC Chemical
Corporation
3676 Davis Rd NW
Dover, OH 44622
NR
2012
Domestic
Manufacturin
g
ND
ND
ND
ND
79,640
NR
2012
Importe
d
CBI
CBI
CBI
CBI
ND
NR
2012
Exporte
d
0
0
0
0
0
NR
201
2
Use
don
Site1
N/A
N/A
N/A
N/A
0
NR
2010 Past
Production
(import and
manufactur
e)
426,480
CBI
CBI
CBI
124,575
NR
2011
National
Productio
n
1 million to
10 million
Withheld
Withheld
NR
^he total volume (domestically manufactured and imported) of the chemical used at the reporting site without leaving the site.
ND = No Data; the company did not provide the requested information.
NR = No Reports; "No Reports" in the CDR public database indicates that production volume was not reported for the IUR/CDR for a
given year. This does not necessarily indicate that a chemical was not manufactured in the United States, but rather indicates that a
chemical, if manufactured, had a production volume below the reporting threshold.
N/A = Not Applicable; the imported chemical was never physically at the site.
"Withheld" in the CDR public database indicates that the national production volume of a chemical was unable to be aggregated in
order to protect to CBI claims.
Source: EPA(2014b)
A-5
Uses
Page 77 of 135
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The 2012 CDR data are minimal for industrial and consumer uses of TBBPA-bis(dibromopropyl ether)
and TBBPA bis(allyl ether), as shown in Table_Apx A-4.
The CDR database does not contain records for TBBPA-bis(methyl ether) (EPA, 2014b). In 1995 the IPCS
Environmental Health Criteria for TBBPA and derivatives stated that as of 1994 TBBPA-bis(methyl
ether) was not used for commercial purposes based on the personal communications from a Great
Lakes Chemicals representative (IPCS, 1995). Instead, it was noted that this compound was a product
of environmental biotransformation.
The Environmental Health Criteria for flame retardants (IPCS, 1997) listed TBBPA-bis(methyl ether) as
a flame retardant for expandable polystyrene, but no references were provided. Furthermore, under
EPA's Design for the Environment Program, companies that produce polystyrene foam did not report
using any TBBPA-based products, including TBBPA-bis(methylether), in flame retardant applications.
Although there is one report of TBBPA-bis(methyl ether)'s use as a flame retardant, multiple other
sources indicate that it is not used in products. Thus, for purposes of this assessment it is assumed that
TBBPA-bis(methyl ether) appears in the environment solely as a transformation product of TBBPA.
Page 78 of 135
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Table_Apx A-4: Industrial and Consumer Use Data for TBBPA-bis(dibromopropyl ether), TBBPA-bis(allyl ether) and TBBPA-bis(methy) ether)
Chemical
Name
TBBPA-
bisfdibromo
propyl ether)
TBBPA-bis(allyl
ether)
TBBPA-
bis( methyl
ether)
Manufacturing Site
ICL-IP America, Inc.
622 Emerson Road, Suite 500
St. Louis, MO 63141-6742
ICC Chemical Corporation
460 Park Avenue
New York, NY 10022- 1906
CBI
ICC Chemical Corporation
460 Park Avenue
New York, NY 10022- 1906
Dover Chemical Corp
3676 Davis Road North West
Dover, OH 44622-9771
NR
Type of
Processing
Processing-
incorporation into
formulation,
mixture or reaction
product
ND
Processing-
incorporation into
article
ND
ND
NR
Industrial Use Data
Sector
Plastics Product
Manufacturing
ND
Plastics Material
and Resin
Manufacturing
ND
ND
NR
Industrial
Use
Flame
retardants
ND
Flame
retardants
ND
ND
NR
Percent of
Production
Volume
100
ND
100
ND
ND
NR
Consumer Use Data
Consumer Use
Product
Category
Plastic and
Rubber Products
not covered
elsewhere
ND
Plastic and
Rubber Products
not covered
elsewhere
ND
ND
NR
Commercial
or Consumer
Use
Commercial
ND
Commercial
ND
ND
NR
Percent of
Production
Volume
100
ND
100
ND
ND
NR
ND = No Data; the company did not provide the requested information.
NR = No Reports; "No Reports" in the CDR public database indicates that production volume was not reported for the IUR/CDR for a given year. This does not necessarily indicate that a chemical was
not manufactured in the United States, but rather indicates that a chemical, if manufactured, had a production volume below the reporting threshold.
Source: EPA(2014b)
Page 79 of 135
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A-6 Fate Properties
If released to air, degradation of these substances by sunlight and reactants in the atmosphere to less
brominated simpler substances is expected to be slow. If released to water, sediment or soil, the fate
of these cluster members may be influenced by partitioning to suspended solids, soil and sediment,
respectively. Microbial biodegradation to less brominated substances can occur in the absence of
oxygen (anaerobic conditions). The biodegradation of these substances in the environment is
dependent on a number of factors including the presence of acclimated microorganisms capable of
biodegrading the chemicals in those media and oxygen levels in the media (anaerobic conditions
promote reductive debromination). A range of degradation rates is possible (from minutes to years).
TBBPA-bis(methyl ether), TBBPA-bis(allyl ether) and TBBPA-bis(dibromopropyl ether) have high
estimated bioaccumulation factors.
Table_Apx A-5 lists fate endpoints that are available for the three cluster members and information is
discussed in subsequent sections.
Table_Apx A-5: Environmental Fate Endpoints for Three Cluster Members
Endpoint
Photo-
degradation
Half-life
Hydrolysis
Half-life
Biodegra-
dation
Bioconcen-
tration
Log Koc
Fugacity3
(Level III
Model)
Air (%)
Water (%)
Soil (%)
Sediment (%)
Persistence0
Bioaccumu-
lation0
TBBPA-bis
(dibromopropyl ether)
1 day (estimated, l.SxlO6
hydroxyl radicals per cm3;
12-hour day)3
Stable
1% after 28 days (not readily
biodegradable, OECD 301C)b
BAF = 1.2xl04 (estimated)3
6.8 (estimated)3
<0.1
4.9
94.8
0.2
P3 (high)
B3 (high)
TBBPA-bis(allyl ether)
1.9 hours (estimated,
l.SxlO6 hydroxyl radicals
per cm3;
12-hour day)3
Stable
No data
BAF = 3.9xl05
(estimated)3
5.8 (estimated)3
<0.1
5.9
93.0
1.1
P3 (high)
B3 (high)
TBBPA-bis(methyl ether)
2.2 days (estimated,
l.SxlO6 hydroxyl radicals
per cm3; 12-hour day)3
Stable
No data
BAF = 8.6xl06
(estimated)3
4.8 (estimated)3
<0.1
4.4
92.3
3.2
P3 (high)
B3 (high)
aEPA(2013a)
b NITE (2010a)
c Criteria specified in: EPA (1999a)
Page 80 of 135
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A-6-1 Fate in Environmental Media
A-6-1-1 Air
TBBPA-bis(methyl ether) and TBBPA-bis(dibromopropyl ether) are expected to undergo relatively slow
atmospheric hydroxy radical oxidation with estimated atmospheric half-lives of a few days. TBBPA-
bis(allyl ether) is expected to have a much shorter half-life (approximately 2 hours) (EPA, 2013a).
However, the relatively low vapor pressures of these substances suggest that they will not exist in the
vapor phase under environmental conditions so atmospheric photo-oxidation is likely to be a slow
process.
A-6-1-2 Fate in Water
Hydrolysis is expected to be relatively unimportant based on the chemical structures of these
compounds. TBBPA-bis(dibromopropyl ether) is not readily biodegradable (NITE, 2010a). Volatilization
from water surfaces will be a relatively unimportant fate process for these compounds based upon an
estimated Henry's law constant of <1.0 x Id'10 atm-m3/mole (EPA, 2013a).
A-6-1-3 Fate in Soil, Sediment and Groundwater
These cluster members are expected to have low mobility in soil based on estimated log Koc values
ranging from 4.8 - 6.8 (EPA, 2013a).
A-6-2 Bioconcentration/Bioaccumulation and Persistence
Persistence and bioconcentration potential are qualitatively characterized according to the criteria set
forth in EPA/OPPT's New Chemicals Program (EPA, 1999a). These cluster members lack experimental
bioconcentration and bioaccumulation data. However, based on the estimated BCF and BAF values for
the substances they are expected to have high bioconcentration potential.
A-7 Exposure
There are a variety of monitoring data for the three cluster members. For the current assessment,
EPA/OPPT consulted data adequacy guidance available for the High Production Volume (HPV) Program
and specific guidance on using exposure data developed by the Organisation for Economic Cooperation
and Development (OECD) (OECD, 2003). Table_Apx A-6 presents a summary table of available
monitoring data, and more detailed information is in Supplemental Files 2 and 3.
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Table_Apx A-6: Availability of Exposure Data for Three Cluster Members
CAS NUMBER
CHEMICAL NAME
21850-44-2
TBBPA-
bis(dibromo
propyl ether)
25327-89-3
TBBPA-
bis(allyl ether)
37853-61-5
TBBPA-
bis(methyl ether)
BIOMONITORING (HUMAN)
Blood
Breast Milk
Adipose Tissue
Placenta
Urine
HUMAN EXPOSURE
Dust ingestion
USGS NWIS DATA
Water
Suspended sediment
Solids
Biota
AIR
Ambient Air
Indoor Air
SOIL
INDOOR DUST
•
•
•
•
SEDIMENT
Freshwater
Marine
SLUDGE
amended soil
biosolids
landfill
sewage
•
•
•
•
•
•
•
'
•
•
WATER
drinking water
groundwater
leachate
precipitation
surface water
wastewater
•
•
•
•
•
•
•
Page 82 of 135
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Table_Apx A-6: Availability of Exposure Data for Three Cluster Members
CAS NUMBER
CHEMICAL NAME
21850-44-2
TBBPA-
bis(dibromo
propyl ether)
25327-89-3
TBBPA-
bis(allyl ether)
37853-61-5
TBBPA-
bis(methyl ether)
AIR AND WATER
deposition
BIOTA
avian
fish
aquatic animals (including
shellfish)
terrestrial animals
vegetation
•
•
•
•
•
•
•
•
•
•
•
•
•
= some data available, US or international
A-8 Ecological Hazard
A-8-1 TBBPA-bis(dibromopropyl ether)
There are a few studies for TBBPA-bis(dibromopropyl ether). Algae, fish and aquatic invertebrates were
exposed to TBBPA-bis(dibromopropyl ether) in different water accommodation fraction (WAF) toxicity
studies (ECHA, 2013). However, the compound was tested at 100 mg/L or higher, which is significantly
above its water solubility (ECHA, 2013). In addition to testing above the water solubility, the
researchers did not measure the test concentrations. As a result, these studies were not acceptable
and were not included in this review.
A 56-day reproduction study is available for the earthworm Eiseniafetida. The study was conducted
using OECD test guideline 207. A chronic value of 724 mg/kg was determined (ECHA, 2013).
No chronic aquatic toxicity studies using fish or aquatic invertebrates (either water column or
sediment) were found forTBBPA-bis(dibromopropyl ether).
A-8-2 TBBPA-bis(allyl ether)
EPA/OPPT did not locate standard ecotoxicity studies that evaluated the effects of TBBPA-bis(allyl
ether) on organisms in the environment.
Page 83 of 135
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A-8-3 TBBPA-bis(methyl ether)
EPA/OPPTdid not locate standard ecotoxicity studies that evaluated the effects of TBBPA-bis(methyl
ether) on organisms in the environment.
A-9 Human Health Hazard
A-9-1 TBBPA-bis(dibromopropyl ether)
EPA/OPPT estimates that TBBPA-bis-(dibromopropyl ether) would not be absorbed through the skin as
a neat material, would have poor skin absorption when in solution and would have poor absorption via
the lungs and gastrointestinal tract.
Following single or repeated (5 or 10 days) oral administrations of 20 mg/kg [14C]-TBBPA-
bis(dibromopropyl ether) to male F-344 rats, the compound was poorly absorbed from the
gastrointestinal tract and uptake to the systemic circulation was considered slow. Distribution to the
tissues accounted for <1% of the dose at 96 hours. Ninety-five percent of the dose was excreted in the
feces within 36 hours of administration after conjugation via glucuronidation and/or sulfation.
Elimination in the urine accounted for <0.1% of the administered dose and 1% of the dose was
excreted in the bile after 24 hours (ECHA, 2013; Knudsen et al., 2007).
Based on oral and dermal LDso values >2, 000 mg/kg and an inhalation LCso value >20 mg/L in rats, the
acute mammalian toxicity of TBBPA-bis(dibromopropyl ether) is considered low (ECHA, 2013). Oral and
dermal LDso values were >20,000 mg/kg in mice (IPCS, 1995). These data indicate that this compound
has low acute toxicity.
Mice administered TBBPA-bis(dibromopropyl ether) in their diet at 200 or 2,000 mg/kg-bw/day for 90
days showed no mortality or abnormal symptoms upon gross pathological examination. The NOAEL
was determined to be 2,000 mg/kg-bw/day (highest dose tested) (ECHA, 2013; IPCS, 1995).
TBBPA-bis(dibromopropyl ether) is not an eye or skin irritant in rabbits. The compound was not a skin
sensitizer in a guinea pig maximization test (ECHA, 2013).
Although the compound was reported to be mutagenic to Salmonella typhimurium strains TA100 and
TA1535 in one assay, it was negative in two other bacterial reverse mutation assays. In the assay that
was positive for genotoxicity, the compound was slightly less pure (95.1 vs. 99.8%) than the batch used
in one of the negative assays that used a very similar method. The test substance purity was not stated
in the third bacterial assay. In an assay using mouse lymphoma cells, the compound was also negative
for mutagenicity (ECHA, 2013; GLCC, 1982). TBBPA-bis(dibromopropyl ether) did not cause
chromosomal aberrations or sister chromatid exchanges in Chinese hamster ovary (CHO) cells, was
Page 84 of 135
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negative in an in vivo micronucleus assay in mice and did not produce unscheduled DNA synthesis in
rats (IPCS, 1995).
The dibromopropyl groups in TBBPA-bis-(dibromopropyl ether) have alkylating potential. However, it is
unlikely that this compound can act as an alkylating agent due to its large molecular weight, low water
solubility and high log Kow. These properties all reduce the bioavailability of the compound. This is
supported by the genotoxicity data of TBBPA-bis(dibromopropyl ether), which are predominantly
negative. TBBPA-bis(dibromopropyl ether) appears to inhibit sulfation of estradiol (E2), but does not
exhibit estrogenic activity via interference with estrogen receptors (ER). TBBPA bis (2,3-dibromopropyl)
ether also does not appear to interfere with AhR-mediated, androgenic or progestrogenic pathways
(Canton et al., 2006). TBBPA-bis(dibromopropyl ether) competed with thyroid hormone precursor
thyroxine (T4) for binding to human transthyretin (TTR), but did not exhibit thyroid hormone (T3)
mimicking activity (Hamers et al., 2006).
A-9-2 TBBPA-bis(allyl ether)
TBBPA-bis(allyl ether) has low acute oral and dermal toxicity. The oral LDso in rats was > 5000 mg/kg
bw, and the dermal LD5o in rabbits was > 2000 mg/kg (Abbott et al., 1981; EC/HC, 2013).
TBBPA-bis(allyl ether) was negative in Salmonella and Saccharomyces when tested with and without
metabolic activation (Brusick, 1977; EC/HC, 2013).
Environmental fractions of TBBPA-bis(allyl ether) induced high cytotoxicity in neuronal cells of primary
cultured cerebellar granule cells. However, the cytotoxic effect was not confirmed by studies in human
liver carcinoma Hep G2, human breast cancer MCF-7 and mouse leukemic monocyte macrophage RAW
264.7 cell lines (EC/HC, 2013; Qu et al., 2011).
A-9-3 TBBPA-bis(metiiyletiier)
No toxicity data were found for TBBPA-bis(methyl ether). Replacement of the hydrogen atom of the
two hydroxy groups in TBBPA by methyl groups slightly increases the size of the compound. There are
no structural alerts for genotoxicity and carcinogenicity of TBBPA-bis(methyl ether). The overall toxicity
of TBBPA-bis(methyl ether) is not expected to be higher than TBBPA based on structure-activity
relationship analysis.
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Appendix B Literature Searches and Data Quality
B-l Literature Available for the Cluster Members
During background, scoping and problem formulation, EPA/OPPT reviewed other recently performed
assessments and searched published and unpublished literature. The literature review included the
cluster members' chemistry, uses, sources (including industrial releases), fate, exposure and hazard to
humans and ecological receptors.
First, EPA/OPPT reviewed previous risk assessments to determine whether various scenarios could
result in risks within the United States. The primary sources of information were human health and
environmental risk assessments for TBBPA conducted by the EU (EC, 2006, 2008) and health and
environmental risk assessments conducted by Canada for TBBPA and TBBPA-bis(allyl ether) (EC/HC,
2013). Other assessments included several predictions of the intake of TBBPA from its presence in food
(de Winter-Sorkina et al., 2003; Driffield et al., 2008; EFSA, 2011; FSAI, 2010; Shi et al., 2009).
In addition, EPA/OPPT reviewed hazard profiles prepared under EPA's Design for the Environment
Program for two members of the TBBPA cluster. The hazard profile for TBBPA is contained within
Flame Retardants in Printed Circuit Boards: Updated Draft Report (EPA, 2014e). A recent profile is
available forTBBPA-bis(2,3-dibromopropyl ether) in the final document An Alternatives Assessment for
the Flame Retardant Decabromodiphenyl Ether (DecaBDE) (EPA, 2014a) and was consulted for the
current TBBPA cluster assessment.
Another source of hazard and exposure information along with physical-chemical and fate properties
for both TBBPA and TBBPA-bis(dibromopropyl ether) is the public data available for chemicals
submitted to the European Chemicals Agency under REACH (ECHA, 2014).
In addition to reviewing these sources, EPA/OPPT conducted a review of toxicology and exposure
studies published through June 2013. EPA/OPPT searched Toxline and Pubmed from the US National
Library of Medicine for toxicology, biomedical and health literature. EPA/OPPT also searched the
Chemical Abstracts Service from the American Chemical Society for chemical information. Other
sources of information reviewed were the publicly available databases such as EPA's Chemical Data
Reporting (CDR) and Inventory Update Reporting (IUR) databases.
EPA/OPPT also searched for additional exposure information in other databases for the time period of
2007 through August 2013. The Web of Science includes multidisciplinary science information from
Thomson Reuters; the BIOSIS Citation Index contains life sciences information and is also from
Thomson Reuters; CAB Abstracts contains life sciences literature from CABI; and Medline contains
biomedical literature from the US National Library of Medicine. Search terms included the chemical
CAS number and chemical abbreviation as well as concentration, environmental, monitoring, human,
exposure, urine, blood, air, water, soil and sediment.
EPA/OPPT's summary of the environmental hazard of TBBPA is based in part on experimental studies
previously summarized in the 2013 Canadian Assessment (EC/HC, 2013) and the final European Risk
Page 86 of 135
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Assessment Report (EC, 2008), study reports from EPA's TSCATS database, public literature searches
and confidential sources. Although all data identified in confidential sources were evaluated, only
information already made public (either in published risk assessments or journal articles) is included in
this risk assessment. In addition, EPA/OPPT searched the ECOTOX database to identify peer-reviewed
articles (EPA, 2014c). Source articles from these searches were retrieved and reviewed.
In some cases (e.g., when a significant pathway was identified), more recent literature (after August or
June 2013) was also consulted.
B-2 Data Adequacy
Data was acceptable if it met standard quality criteria, which varies according to the type of
information reviewed. Most toxicity or exposure monitoring studies referenced in previous risk
assessments, such as those conducted by the European Union (EU) or Canada, were considered to be
adequate if the previous assessment indicated that the studies were valid. However, EPA/OPPT
reviewed selected studies referenced in such assessments when either the interpretation and
description of results or the adequacy of the study was not clearly stated.
For the studies described in previous assessments that EPA/OPPT evaluated further and for recently
published studies, EPA/OPPT followed data adequacy guidance developed for the High Production
Volume (HPV) Chemicals Program (EPA, 1999b). The guidance generally suggests that EPA use studies
conducted according to accepted testing guidelines such as EPA harmonized guidelines for pesticides
and industrial chemicals or Organisation for Economic Cooperation and Development (OECD) test
guidelines. Scientific quality criteria established for the HPV Program include: a clear description of the
endpoints, inclusion of appropriate controls, identification of test substance and test organism, stated
exposure duration and administration route, transparent reporting of effect concentrations and
adherence to recommended tests strategies (EPA, 1999b). These criteria are based on guidelines
developed and used by EPA's Office of Chemical Safety and Pollution Prevention and OECD (EPA,
2014g; OECD, 2015). The guidance also allows EPA/OPPT to use data that don't strictly follow such
guidelines, primarily when less acceptable data can support the conclusions from a larger body of
studies for a particular endpoint using a weight-of-evidence approach.
EPA/OPPT also used study reliability criteria presented in a 2003 OECD Guidance Document on
Reporting Summary Information on Environmental, Occupational and Consumer Exposure (OECD, 2003)
to evaluate studies that reported environmental and residential (indoor dust, etc.) monitoring data and
biomonitoring studies.
Finally, EPA/OPPT used scientific judgment based on an understanding of TBBPA's unique
physicochemical and fate characteristics to judge data quality and whether a particular study can
contribute to TBBPA's overall hazard identification. This scientific evaluation allowed EPA/OPPT to
consider whether any identified study deficiencies would detract from the results or would be minor
enough so that EPA/OPPT could still rely on the study conclusions for hazard identification.
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Appendix C Fate and Transport
The environmental transport and transformation of TBBPA is strongly influenced by its low volatility,
low water solubility and its slow degradation by biotic and abiotic processes in the environment.
Where particulate TBBPA is released to air its oxidation by OH radicals is expected to be slow. TBBPA
may be subject to photolysis in air but the rates and products are unknown. Ultimately air releases of
TBBPA would be expected to undergo deposition to terrestrial and aquatic environments. TBBPA tends
to partition to soil and sediment where it can undergo microbial 0-methylation to form TBBPA-
bis(methyl ether). It can also undergo debromination under anaerobic conditions to form bisphenol A
(BPA). Although these transformations have been observed, the overall environmental persistence of
TBBPA is expected to be moderate to high. Measured bioconcentration factors indicate that TBBPA has
a low potential for bioconcentration.
The rate of loss of the substance, typically due to degradation, may be less than its rate of entry to the
environment. Therefore, levels of the compound in the environment may increase overtime, leading
to greater potential for exposure. EPA/OPPT will assume that loss is slow and that levels will increase
over time when estimating exposure in the proposed assessment.
Table_Apx C-l summarizes environmental fate data for TBBPA.
Table_Apx C-l: Environmental Fate
Endpoint
Results
Photodegradation
half-life
3.6 days (estimated, l.SxlO6 hydroxyl radicals per cm3; 12-hour day)3
Hydrolysis Half-life
Stable
Biodegradation
18.1 - 25.7% after 64 days (Massachusetts sandy loam, aerobic conditions)13;
59.9 - 64.1% after 64 days (Arkansas silty loam, aerobic conditions)13;
56.8 - 58.9% after 64 days (California clay loam, aerobic conditions)13;
43 - 56.3% after 64 days (Massachusetts sandy loam, anaerobic conditions)0;
35 -46.6% after 64 days (Arkansas silty loam, anaerobic conditions)0;
9.4 - 10.5% after 64 days (California clay loam, anaerobic conditions)0;
0% after 14 days (not readily biodegradable, OECD 301C)d
Bioconcentration
BCF = 720 (measured in eastern oysters)6
BCF = 1,200-1,300 (measured in fathead minnows)11;
BCF = 30 - 341 (measured in carp based on a water concentration of 0.08
mg/L)d;
BCF = 52 - 485 (measured in carp based on a water concentration of 0.008
mg/L)d;
BCF = 20 - 170 (measured in bluegill for water concentrations of 0.0098 -
0.0014 mg/L)f
BAF = 717.5 (estimated)3
Log Koc
5.4 (estimated)3
4.8 (measured)8
6.0 (measured)8
5.2 (measured)8
5.0 (measured)8
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Table_Apx C-l: Environmental Fate
Endpoint
Fugacity3
(Level III Model)
Air (%)
Water (%)
Soil (%)
Sediment (%)
Results
<0.1
1.8
69.1
29.1
a EPA (2013a);b GLCC (1989c);c GLCC (1989d); dNITE (2010b);e GLCC (1989b);f Stoner_Laboratories (1978);
' GLCC (1989e);h GLCC (1989a)
C-l Fate in Environmental Media
C-l-1 Fate in Air
TBBPA is expected to undergo relatively slow atmospheric hydroxy radical oxidation. Estimated
atmospheric half-lives of a few days were predicted using Version 4.10 of EPISuite (EPA,
2013a). However, the relatively low vapor pressure of TBBPA suggests that it will not exist in the vapor
phase under environmental conditions. Therefore, atmospheric photo-oxidation is likely to be a slow
process. TBBPA has been shown to undergo direct photolysis in water when irradiated with UV light in
the environmental spectrum (300 - 390 nm) and may be subject to photolysis in the atmosphere
(Eriksson etal., 2004).
C-l-2 Fate in Water
TBBPA released to water will sorb to suspended solids and sediments due to its low water solubility
and high log Koc value. Hydrolysis of TBBPA is expected to be relatively unimportant due to the
absence of structural components that hydrolyze under environmental conditions. Volatilization from
water surfaces will be a relatively unimportant fate process based upon an estimated Henry's law
constant of <1.0 x 10'10 atm-m3/mole (Thomas, 1982b).
In studies of photolysis, TBBPA was irradiated in water with the range of solar UV wavelengths that are
encountered in the environment. Photolysis half-lives ranged from 16 minutes at pH 10 to 350 minutes
at pH 5.5. TBBPA was shown to photodegrade via cleavage between the tertiary carbon and one of the
benzene rings. The main decomposition products were 4-(2- hydroxyisopropyl)-2,6-dibromophenol; 4-
isopropylene-2,6-dibromophenol; and 2,6-dibromo-4-isopropylene (Eriksson et al., 2004).
TBBPA is not readily biodegradable. It achieved 0% of its theoretical oxygen demand over a 14-day
incubation period using an activated sludge inoculum and the modified MITI (OECD 301C) test (NITE,
2010b). In an aerobic river water and sediment biodegradation study TBBPA degraded with half lives in
the range of 48-84 days (Fackler, 1989).
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C-l-3 Fate in Soil and Sediment
TBBPA is expected to have low mobility in soil based on both estimated and measured log Koc values.
Larsen et al. (2001) investigated the leaching potential of TBBPA applied to soil and sand columns and
found no TBBPA leached from the bottom of the soil column with much of it retained in the upper 1 cm
of the soil. The adsorption of TBBPA to sediment was determined during a toxicity study on sediment-
dwelling midge larvae (GLCC, 1989e). The solid-phase concentrations were not measured during the
study and are based on the nominal amount of substance added. The mean value for the organic
carbon-water partition coefficient (Koc) that was estimated from the data was 68,753 I/kg (log Koc 4.8).
In addition, as part of the 14-day toxicity study, sediment-solids and sediment pore water
concentrations were measured allowing estimates of the value of Koc to be made. The mean values of
Koc obtained were 1,008,730,141,980 and 94,830 I/kg (log Koc 6.0, 5.2 and 5.0) for sediments with
organic carbon contents of 0.25, 2.7 and 6.8% respectively.
Hydrolysis of TBBPA in soil and sediments is expected to be relatively unimportant due to the absence
of structural components that hydrolyze under environmental conditions. Volatilization of TBBPA from
moist soil surfaces is not expected to be an important fate process given its low Henry's law constant
(<1.0 x 10"10 atm-m3/mole) (Thomas, 1982a). TBBPA is not expected to volatilize from dry soil surfaces
based upon its low vapor pressure.
The biodegradation potential of TBBPA under aerobic and anaerobic conditions in a 64-day test was
investigated using three soils. Under aerobic conditions, TBBPA was degraded 18- 64% (GLCC, 1989c).
Under anaerobic conditions, the degradation range was 9 - 56 % (GLCC, 1989d). Two main
biodegradation products were noticed in each soil. However, these were not positively identified and
did not appear to be the dimethyl derivative TBBPA-bis(methyl ether) or the diethyl derivative (GLCC,
1989c, 1989d). In another study, biodegradation half-lives of 65 days, 93 days and 430 days were
reported for TBPPA in an aerobic activated sludge, aerobic digested sludge and anaerobic activated
sludge, respectively (Nyholm et al., 2010).
TBBPA has been shown to degrade to BPA under anaerobic conditions and form the intermediates
tribromobisphenol A (Tri BBPA), dibromobisphenol A (DiBBPA) and monobromobisphenol A, which
were rapidly consumed (Arbeli and Ronen, 2003). TBBPA has also been demonstrated to undergo
microbial 0-methylation by bacterial isolates and sediments to form the TBBPA derivative, TBBPA-
bis(methyl ether) (Allard et al., 1987). Conversions to the methylated derivative of approximately 60%
in 24 hours were observed with pure culture bacterial isolates. In sediments, conversion rates ranged
from 10% in 60 days to 50% in 80 days (George and Haggblom, 2008).
C-2 Persistence
Biotic and abiotic degradation studies have shown TBBPA to degrade very slowly under most
environmental conditions with half-lives greater than 2 months in water, soil and sediment (Fackler,
1989; GLCC, 1989c, 1989d; NITE, 2010b).
The persistence of TBBPA is influenced by a number of factors. These include the environmental media
to which the substance partitions; the presence of acclimated microorganisms capable of biodegrading
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the chemical in those media; oxygen levels in the media (anaerobic conditions promote degradation of
TBBPA by reductive debromination) and prolonged direct exposure to sunlight leading to photolysis of
the molecule. When microorganisms are not acclimated, oxygen levels are high and direct exposure to
sunlight is limited, TBBPA has been shown to be resistant to environmental degradation processes.
C-3 Bioaccumulation/Bioconcentration
The majority of bioconcentration studies in fish and mollusks indicate that TBBPA has a low
bioconcentration potential. However, a single study using fathead minnows reported a BCF value of
1200-1300 (GLCC, 1989a). In this study that used 14C-TBBPA, radioactivity was eliminated from the
tissues with a half life of less than 24 hours. In further analysis of the data, as discussed in EC (2008),
the BCF was calculated in terms of a ratio of parent TBBPA to the total body burden of radioactivity,
resulting in a BCF of approximately 160 for just TBBPA and not its degradates. This recalculated value is
consistent with the results of other fish bioconcentration studies.
Based upon the experimental evidence and expert opinion, TBBPA is expected to have low
bioaccumulation potential.
C-4 Exclusion of Degradation Products from Further Assessment
For a full consideration of possible risks from the manufacture, use and disposal of TBBPA, EPA/OPPT
identified the most likely compounds that could result from combustion, biodegradation or photolysis
of TBBPA. Based on data available for these compounds, EPA/OPPT determined whether the
compounds might result in risks to human health or the environment and whether it would be feasible
to assess risks from these compounds in the current assessment. As a result of this scoping analysis,
EPA/OPPT concluded that data on degradation are limited, uncertain or both. Therefore, EPA/OPPT will
not assess risks from TBBPA's degradation products in a risk assessment.
C-4-1 Combustion Products
Incineration of TBBPA can result in polybrominated dibenzo-p-dioxins (PBDDs) and polybrominated
dibenzofurans (PBDFs) as well as polyaromatic hydrocarbons (PAHs).
A recent study found that PBDDs, PBDFs and PAHs were emitted from incineration of TBBPA epoxy
laminates. PAHs were emitted at higher levels from this laminate than from non-flame retardant
laminates (Sidhu et al., 2013). In another study, Wichmann et al. (2002) found that PBDDs and PBDFs
were emitted at similar magnitudes when comparing emissions from TBBPA used in reactive
applications to those in additive flame retardant applications with PBDFs released in higher amounts
that PBDDs.
An accurate estimate of the amount of TBBPA within electronic waste is not available. Furthermore,
EPA/OPPT doesn't have robust information on the amount of electronic waste that is incinerated in the
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United States. Finally, compounds other than TBBPA can result in similar combustion products when
incinerated. Therefore, the contribution of TBBPA to combustion byproducts is not possible to
determine with enough accuracy to include in EPA/OPPT's proposed risk assessment.
C-4-2 Biodegradation Products
Bisphenol A (BPA; CAS 80-05-7) is a possible product of reductive debromination of TBBPA, primarily
under anaerobic conditions. Thus TBBPA could be a source of BPA in the environment (EC, 2008).
Overall, biodegradation data are considered to be too limited to predict, with confidence, the rate at
which TBBPA degrades to BPA in the environment. This is because the majority of the studies use
microorganisms that have been collected from environments contaminated with TBBPA, exposed to
TBBPA over extended periods to induce adaptation to degrade the substance and are conducted under
laboratory conditions that are not necessarily representative of the environment.
C-4-3 Photodegradation Products
TBBPA may photodegrade to form a range of bromophenols and dibromoisopropylphenol derivatives.
Overall there appears to be limited to no human health toxicity data for dibromophenols. Some
ecotoxicity data are available for 2-bromophenol (CAS 95-56-7) (NLM, 2009). Due to uncertainties in
extrapolation from laboratory to the field, it is not certain how much of these products would be
formed in the environment (Eriksson et al., 2004).
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Appendix D Toxics Release Inventory Emissions
This appendix describes EPA/OPPT's approach to assessing exposure surrounding manufacturing
facilities. For both the environment and human health, EPA/OPPT modeled the impact of TBBPA
deposition from TRI-reported air releases from two manufacturing facilities that emitted the highest
amounts of TBBPA.
D-l TRI Releases from Manufacturers
In this assessment, EPA/OPPT used data from the Toxics Release Inventory (TRI) to characterize
environmental releases of TBBPA. The TRI is a database that contains detailed information on releases
and transfers of certain listed toxic chemicals from industrial facilities. The database is maintained by
the Agency and is updated annually (EPA, 2012e).
In 2012, a total of 52 facilities across 7 industries reported releases of TBBPA to TRI (EPA, 2012e). This
information is reported in Table_Apx D-l. See Table_Apx D-2 for a more detailed account of these
disposals and releases.
Based on data from the 2012 TRI, the chemicals industry accounted for ~97% of all reported TBBPA
disposal and releases. Two facilities from the chemicals industry accounted for ~94% of these disposal
and releases. Of the total disposal and releases, 94.5% were to land and 4.5% to air.
Table_Apx D-l: Disposal and Releases of TBBPA by Industry as Reported in the 2012 TRI
Industry
Chemicals
Textiles
Hazardous
Waste/Solvent
Recovery
Transportation
Equipment
Plastics and Rubber
Paper
Computers/Electronics
Products
NAICS3
Code
325
313/314
562
336
326
322
334
Total
Number
of
Facilities
23
5
4
8
6
1
5
52
Disposal and Releases (pounds per year)
On-site
75,786
143
662
8
33
5
1
76,637
Off-Site
48,174
1,557
762
560
154
0
1
51,208
Total
123,960
1,700
1,425
567
187
5
2
127,845
Percent of
Total
97.0%
1.3%
1.1%
0.4%
0.1%
<0.1%
<0.1%
100%
a - North American Industry Classification System (NAICS)
Source: EPA (2012e)
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Table_Apx D-2: Disposal and Releases of TBBPA by Industry as reported in the 2012 TRI
Industry
Chemicals
Textiles
Hazardous
Waste/Solvent
Recovery
Transportation
Equipment
Plastics and
Rubber
Paper
Computers/
Electronics
Products
NAICS
Code
325
313/314
562
336
326
322
334
Total
Total On-site Disposal and Releases
Under-
ground
Injection
Class 1
Wells
100
0
0
0
0
0
0
100
RCRA
Subtitle C
Landfills
0
0
656
0
0
0
0
656
Other
On-Site
Landfills
70,182
0
0
0
0
0
0
70,182
Fugitive
Air
Emissions
44
141
0
1
33
5
0
224
Stack Air
Emissions
5,453
2
7
6
0
0
1
5,469
Surface
Water
Discharges
6
0
0
0
0
0
0
6
Total Off-site Disposal and
Releases
RCRA
SubtitleC
Landfills
44,205
0
0
0
0
0
0
44,205
Other
Landfills
3,969
1,557
762
29
154
0
1
6,471
Other
Land
Disposal
0
0
0
531
0
0
0
531
Total On-
and Off-site
Disposal and
Releases
123,960
1,700
1,425
567
187
5
2
127,845
Source: EPA(2012e)
-------
The two top-releasing facilities in the 2012 TRI are shown in Table_Apx D-3. In 2012, the Dow Chemical
facility in Freeport, Texas was one of the two highest emitters. However, historical emissions have
been highest from Great Lakes Chemical (El Dorado, Arkansas) and Albemarle (Magnolia, Arkansas), as
indicated in Table_Apx D-4.
Table_Apx D-3: Disposal and Releases of TBBPA by Manufacturing Facility as Reported in the 2012
TRI
Facility
Dow Chemical
(Freeport, TX)
Albemarle
(Magnolia, AR)
Remaining 50 facilities
NAICS
Code
325
325
N/A
Total
Disposal and Releases (pounds per year)
On-site
68,882
6,410
1,345
76,637
Off-Site
0
44,176
7,032
51,208
Total
68,882
50,586
8,377
127,845
Percent of Total
53.9%
39.6%
6.6%
100%
Source: EPA(2012e)
Over the past 13 years, disposal and releases of TBBPA, as reported in TRI, have decreased by ~84%.
During this time, two facilities from the chemicals industry were, on average, responsible for ~90% of
all TBBPA disposal and releases (see Table_Apx D-4). These disposal and releases were primarily to land
or air.
Table_Apx D-4: TBBPA TRI Release Trends from 2001 to 2012
Year
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
Disposal and Releases from All Facilities
(pounds per year)
On-site
257,134
250,689
190,326
145,766
113,779
106,737
114,902
83,124
48,962
46,906
64,801
58,837
76,637
Off-site
537,846
625,651
863,971
497,653
466,078
253,070
210,464
261,016
153,536
113,089
180,459
157,583
51,208
Total
794,981
876,340
1,054,297
643,419
579,857
359,807
325,367
344,139
202,497
159,995
245,260
216,420
127,845
Disposal and Releases
from Great Lakes
Chemical and Albemarle
Percent of Total
94%
96%
98%
98%
98%
98%
92%
98%
95%
95%
96%
88%
40%
Source: EPA (2012e)
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The stack air emissions data for the two facilities modeled in this assessment are outlined in Table_Apx
D-5. On average, from 2000 to 2012, these two manufacturing plants have accounted for 97% of all
TBBPA emitted to air in the United States. The Great Lakes facility produced the majority of these
emissions from 2000 to 2011. In 2012, this facility reported no emissions to air (EPA, 2012e).
Table_Apx D-5: Air Emissions for Facilities Modeled in the Current Assessment
Year
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
Stack Air Emissions
(pounds per year)
Great Lakes Chemical
41,824
40,455
45,030
61,724
59,989
52,933
55,900
49,156
38,724
40,821
48,692
32,987
0
Albemarle
15,403
11,000
6,000
6,000
6,042
5,982
4,938
5,557
1,684
1,689
1,590
1,590
5,000
Source: EPA(2012e)
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Appendix E TBBPA and other Dust Particle Concentrations in
Occupational Settings
Workers may be exposed primarily by ingesting TBBPA dust from the air at work sites. Limited data are
available for TBBPA at manufacturing and processing sites. These data are described below along with
data for particles not otherwise regulates (PNOR) at the same types of facilities that would
manufacture and process TBBPA. EPA/OPPT will consider all of these data to estimate risks to workers.
Manufacturing
During bagging or loading operations, dust may be generated within facilities that manufacture TBBPA
and workers may be exposed by ingesting particles that are inhaled from the air. To consider potential
workplace exposures within facilities, EPA/OPPT examined Chemical Exposure Health Data (CEHD)
from the Occupational Safety and Health Administration (OSHA, 2015). Because TBBPA is not
specifically listed in OSHA's database, data for particulates not otherwise regulated (PNOR) were used
to develop estimates of TBBPA exposures. This exposure data represents the NAICS codes listed below,
which were reported in the 2012 TRI and are judged to be representative of facilities that accounted
for approximately 97% of TBBPA releases reported in the 2012 TRI:
• 3251 Basic Chemical Manufacturing
• 3252 Resin, Synthetic Rubber and Artificial Synthetic Fibers and Filaments Manufacturing
• 3255 Paint, Coating and Adhesive Manufacturing
• 3259 Other Chemical Product and Preparation Manufacturing
Potential dust exposures related to TBBPA manufacturing (NAICS Code 3251) are summarized in
Table_Apx E-l. EPA/OPPT will evaluate risks of ingesting dust after inhalation by workers inside these
types of facilities.
Table_Apx E-l: Potential Occupational Exposures from Manufacture of TBBPA*
Exposure Type
Potential Dust Exposures
(mg/m3)**
Comments
PNOR
(Total Dust)
Range:
0.47 to 195
30 data points; personal samples; ten years of
data (2002 to 2011); not specific to TBBPA
OSHA PEL for PNOR (Total Dust): Time-
Weighted Average (TWA) 15 mg/m3
NAICS Code: 3251
PNOR
(Respirable
Fraction)***
Range:
0.07 to 19
21 data points; personal samples; ten years of
data (2002 to 2011); not specific to TBBPA
OSHA PEL for PNOR (Respirable
Fraction): TWA 5 mg/m3
NAICS Code: 3251
Source: OSHA (2015)
* Dust concentrations are used as surrogates for TBBPA from similar industry sectors as TBBPA sectors.
**These exposure estimates are not TWA values; thus they cannot be compared directly to the OSHA PELs.
*** < 10 urn based on particle sampling methods (see:
https://www.osha.gov/dsg/topics/silicacrystalline/dust/chapter l.html)
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E-l Processing
During loading and unloading operations, dust may be generated within facilities that process or
compound TBBPA. Dust generated from unloading operations is expected to be pure TBBPA. However,
dust generated from compounding and loading operations is not expected to be pure TBBPPA. These
two types of dust (pure vs. mixture) may have different bioavailabilities. Potential exposures from
processing TBBPA are summarized in Table_Apx E-2. EPA/OPPT will evaluate occupational risk from
exposure in processing plants to TBBPA as/in dust after ingestion from air.
Table_Apx E-2: Potential exposures from processing of TBBPA*
Data Source
EU TBBPA RA
(EC, 2006)
ACC BFR End
User Survey
(ACC, 2000)
OSHA CEHD**
(OSHA, 2015)
Potential TBBPA Exposures
(mg/m3)
0.2 to 12
2 to 50
0.1 to 15
0.1 to 2
0.0007 to 79
0.02 to 268
0.05 to 169
• Personal Samples (n=13)
• Model estimates based on EASE
(Estimation and Assessment of Substance
Exposure)
• Total dust, personal samples (n=52)
• Total dust, personal samples (n=ll); BFR
identified as TBBPA
• Bromine content; personal samples (n=34)
• Total dust; Personal Samples (n=146)
• Respirable dust; Personal Samples (n=63)
Comments
• Monitoring data is for
compounding operations;
represents unloading of
powder; specific to TBBPA.
• Monitoring data represents
BFR end user sites.
• Most measurements are for
total dust samples (not TBBPA
specific).*
• Ten years of data (2002 to
2011)
• Targeted to represent
processing facilities (NAICS
Codes: 3252, 3255, 3259)
• Data is not specific to TBBPA.*
* Data are for dust concentrations that could be used as a surrogate for TBBPA concentrations based on similarities in
industry sectors where the dust is generated.
**These exposure estimates are not TWA values; thus they cannot be compared directly to the OSHA PELs.
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Appendix F Ecological Hazard Study Summaries
The following sections describe standard toxicity studies that were considered for use in the proposed
TBBPA risk assessment for environmental organisms.
The test species, test conditions and toxicity endpoints were summarized and evaluated for data
quality. Data quality inclusion criteria included: use of appropriate analytical and test controls,
identification of test substance and test organism, stated exposure duration, a clear description of the
effect endpoints and transparent reporting of effect concentrations. Guideline studies as well as
studies using other protocols were included if they met data quality criteria.
Available toxicity information on representative test species are used to denote toxicity to a wider
group of organisms within individual taxa or species groups such as aquatic plants, aquatic
invertebrates (both water column and sediment), fish and soil invertebrates.
Because only limited experimental data are available to characterize toxicity to these species and taxa,
acute and chronic uncertainty factors that account for both differences in species sensitivities and
variability among laboratories that conducted the studies can be applied to toxicity values to calculate
lower bound levels on the concentration associated with toxicity to organisms in the species or taxa
described above. These lower bound values are referred to as concentrations of concern (COCs).
The uncertainty factors are based on established EPA/OPPT methods (EPA, 2012d, 2013b) These
factors are dependent upon the availability of datasets that can be used to characterize relative
sensitivities across multiple species within a given taxa or species group, but are often standardized in
risk assessments conducted under TSCA because the data available for most industrial chemicals is
limited (Ahlers et al., 2008).
F-l Toxicity to Aquatic Organisms
Studies considered to be of sufficient quality are summarized in Table_Apx F-l. Some effects were
observed in other acute aquatic studies, but those values were considerably above the reported water
solubility for TBBPA. EPA found a few acceptable studies to characterize chronic toxicity to aquatic
organisms.
F-l-1 Aquatic Plant Toxicity
Toxicity was observed, with 72-hr ECso values ranging from 0.09 to >5.6 mg/L, when algae were
exposed to TBBPA (GLCC, 1988b; Walsh et al., 1987).
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F-l-2 Aquatic Invertebrate Toxicity
F-1-2-1 Abater Column
Acute exposure to TBBPA by freshwater aquatic invertebrates (Daphnia magna) resulted in 48-hour
EC5o values15 of 0.60 and 0.96 mg/L (GLCC, 1978c; Waaijers et al., 2013). In addition, TBBPA was toxic
to several marine invertebrates at low concentrations. The Eastern oyster (Crassostrea virginica)
exhibited a 96-hour ECso value of 0.098 mg/L based on shell deposition (GLCC, 1989b). In an acute
mysid shrimp (Mysidopsis bahia) marine ecotoxicity study, adverse effects from TBBPA were observed
with 96-hour ECso values ranging from 0.86 to 1.2 mg/L (Goodman et al., 1988). A TBBPA concentration
of 0.40 mg/L caused lethality in the marine invertebrate species Acartia tonsa under static conditions
(Wollenberger et al., 2002). In addition, Chironomus tentans were exposed to TBBPA in a 14-day
toxicity study (GLCC, 1989e). Adverse effects (lethality) were observed at 0.13 mg/L (14-day LC5o).
Adverse effects from TBBPA have been observed in chronic invertebrate studies at concentrations of
less than 1 mg/L. Chronic exposure of aquatic invertebrates to TBBPA resulted in a 21-day MATC of
0.540 mg/L based on reproductive effects (SLS, 1989b). In addition, shell length reduction in the
common mussel (Mytilus edulis) was observed in a 70-day GLP study, with a MATC of 0.023 mg/L (ACC,
2005).
F-l-2-2 Sedyjnent
The available sediment toxicity studies for sediment-dwelling worms, emergent insects and amphipods
indicate that observed effects exhibited sediment toxicity in the range of 117 - 500 mg/kg for the
species tested.
Wildlife International (Krueger, 2002a, 2002b) studied survival and reproduction in a repeated 28-day
study using the blackworm Lumbriculus variegates, exposed to TBBPA. For total organic carbon (TOC)
contents of 2 and 5%, the MATCs were 117 and 329 mg/kg sediment dry weight, respectively.
In another 28-day study using the midge Chironomus riparius exposed to TBBPA, a MATC for
emergence was determined to be 177 mg/kg. In this same study, a 28-day ECso of 235 mg/kg dry
weight was derived based on midge emergence (ACC-BFRIP, 2005b). In another sediment toxicity test,
amphipods (Hyalella azteca) were exposed to TBBPA, resulting in a 28-day NOEC and LOEC of 250 and
500 mg/kg dry weight, respectively, based on the survival endpoint (ACC-BFRIP, 2006). The MATC for
H. azteca was 354 mg/kg dry weight.
15 Similar to the LCso, the ECso is the effect concentration at which 50% of the test organisms show a specific effect other
than lethality. The value is calculated using the % of organisms affected at the concentrations used in the toxicity study.
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F-l-3 Fish Toxicity
For a variety offish species, acute toxicity studies resulted in 96-hour LCso values16 ranging from 0. 40
to 1.1 mg/L under static and flow-through conditions with measured and nominal concentrations
(GLCC, 1978b, 1988a).
Adverse effects from TBBPA have been observed in chronic fish studies at concentrations of less than 1
mg/L. A 35-day fish maximum acceptable toxicity concentration (MATC)17 for Pimephales promelas
was 0.22 mg/L for hatching (SLS, 1989c). A zebrafish embryo toxicity test was conducted with TBBPA,
but considered unacceptable for the purposes of this hazard assessment due to a lack of replicates and
difficulties with study interpretation (only body burden concentrations presented) (Kuiper et al., 2007).
Chronic exposure of aquatic invertebrates to TBBPA resulted in a 21-day MATC of 0.540 mg/L (SLS,
1989b). In addition, shell length reduction in the common mussel (Mytilus edulis) was observed in a 70-
day GLP study, with a MATC of 0.023 mg/L (ACC, 2005).
F-l-4 Concentrations of Concern
The most sensitive species is expected to be protective of a wider variety of species that are not
specifically represented by the available experimental data. These were used for both water column
and sediment species, with application of uncertainty factors, to calculate acute and chronic COCs.
F-l-4-1 Water Column
Considering data from acute studies offish, aquatic invertebrates and algae, the C. virginica 96-hour
EC5o of 0.098 mg/L resulted in the lowest value and EPA/OPPT divided this 96-hour EC5o value by an
uncertainty factor (UF) of 5 for acute tests using invertebrates, as per established EPA/OPPT methods
(EPA, 2012d, 2013b) to give an acute COC of 0.02 mg/L (or 20 ug/L).
Although a value of 0.090 mg/L using Skeletonema costatum is the single most sensitive acute toxicity
value, the resulting value (0.0225 mg/L) is not the most sensitive COC when divided by the standard
aquatic plant uncertainty factor of 4.
For chronic concerns, the Mytilus edulis 70-day chronic value of 0.023 mg/L based on growth rate is the
most sensitive value. When divided by an uncertainty factor (UF) of 10 for chronic effects, as per
established EPA/OPPT methods (EPA, 2012d, 2013b), the resulting chronic COC is 0.002 mg/L (or 2
ug/L).
F-l-4-2 Sediment
The L variegates 28-day MATC of 117 mg/L, based on reproduction, was divided by an uncertainty
factor (UF) of 10 for chronic effects, as per established EPA/OPPT methods (EPA, 2012d, 2013b), to give
a chronic COC of 11.7 mg/L (or 11,700 u.g/L).
16 LCso is the lethal concentration at which 50% of the test organisms die. It is calculated using the % of organisms that die at
the concentrations in the toxicity study and is not necessarily a concentration that was used in the study.
17 The MATC is calculated as the geometric mean of the NOEC and LOEC from the study.
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Table_Apx F-l: Toxicity from TBBPA to Aquatic Organisms
Test Organism
Fresh/
Salt
Water
Test Guideline/
Study Type
Duration
Endpoint
Concentration
(mg/L)
Chemical
Analysis
Effects
Reference
Aquatic Plants
Green algae
Pseud o-
kirchneriella
subcapitata
Green algae
Chlorella sp.
Diatom
Skeletonema
costatum
Diatom
Thalassiosira
pseudonana
Fresh
Salt
Salt
Salt
USEPA40CFR
797.1050
USEPA40CFR
797.1050
USEPA40CFR
797.1050
USEPA40CFR
797.1050
72-hr
96-hr
72-hr
72-hr
EC50
EC50
EC50
EC50
>5.6
>1.5
0.09
0.13
Measured
Static
Measured
Static
Measured
Static
Measured
Static
Biomass
Growth
Growth
Growth
GLCC(1988b)
Walsh etal.
(1987)
Walsh etal.
(1987)
Walsh etal.
(1987)
Aquatic Invertebrates: Water Column
Water flea
Daphnia magna
Water flea
Daphnia magna
Eastern oyster
Crassostrea
virginica
Mysid shrimp
Mysidopsis bahia
Marine copepod
Acartia tonsa
Midge
Chironomus
ten tans
Fresh
Fresh
Salt
Salt
Salt
Fresh
OECDTG 202
OECD TG 202,
2004
USEPACFR40TG
797.1800
_a
Draft ISO/DIS
14669
ASTM, 1987
E1706-05
48-hr
48-hr
96-hr
96-hr
48-hr
14-day
LC50
LC50
EC50
LC50
LC50
LC50
0.96
0.60
0.098
0.86
0.40
0.13
Nominal
Static
Nominal
Static
Measured
Flow- through
Measured
Flow-through
Nominal
Semi-static
Measured
Flow-through
Mortality
Mortality
Shell deposition
Mortality
Mortality
Mortality
GLCC(1978c)
Waaijers et al.
(2013)
SLS(1989a)
Goodman et al.
(1988)
Wollenberger et
al. (2002)
GLCC (1989e)
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Table_Apx F-l: Toxicity from TBBPA to Aquatic Organisms
Test Organism
Water flea
Daphnia magna
Blue mussel
Mytilus edulis
Fresh/
Salt
Water
Fresh
Salt
Test Guideline/
Study Type
USEPACFR40
797.1330
_a
Duration
21-day
70-day
Endpoint
LOEC
NOEC
MATC
LOEC
NOEC
MATC
Concentration
(mg/L)
0.98
0.30
0.54
0.032
0.017
0.023
Chemical
Analysis
Flow-through
Measured
Flow-through
Measured
Effects
Reproduction
Growth rate,
shell length
Reference
SLS (1989b)
ACC (2005)
Aquatic Invertebrates: Sediment
Blackworm
Lumbriculus
variegatus
Blackworm
Lumbriculus
variegatus
Harlequin fly
Chironomus
riparius
Amphipod
crustacean
Hyalella azteca
Fresh
Fresh
Fresh
Fresh
USEPAOPPTS
850.1735, ASTM
E1706/95b
USEPAOPPTS
850.1735, ASTM
E1706/95b
OECDTG218
USEPAOPPTS
850.1735
28-day
Tested in 2.5%
OC content
28-day
Tested in 5.9%
OC content
28-day
28-day
LOEC
NOEC
MATC
LOEC
NOEC
MATC
LOEC
NOEC
MATC
LOEC
NOEC
MATC
151
90
117
426
254
329
250
125
177
500
250
354
Measured
Measured
Nominal
Nominal
Reproduction
Reproduction
Emergence
development
Survival
Krueger (2002a)
Krueger (2002b)
ACC-BFRIP (2005b)
ACC-BFRIP (2006)
Fish
Rainbow trout
Oncorhynchus
mykiss
Bluegill sunfish
Lepomis
macrochirus
Fathead minnow
Pimephales
promelas
Fresh
Fresh
Fresh
OECD TG 203,
1984
OECD TG 203,
ASTM, 1975
OECD TG 203,
1984
96-hr
96-hr
96-hr
LC50
LC50
LC50
0.40
0.51
0.54
Nominal
Nominal
Measured
Flow-through
Mortality
Mortality
Mortality
GLCC(1978b)
GLCC (1978a)
GLCC (1988a)
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Table_Apx F-l: Toxicity from TBBPA to Aquatic Organisms
Test Organism
Rainbow trout
Oncorhynchus
mykiss
Zebra fish
Danio rerio
Fathead
minnow
Pimephales
promelas
Fresh/
Salt
Water
Fresh
Fresh
Fresh
Test Guideline/
Study Type
OECD TG 203
OECDTG202
USEPACFR40
797.1600
Duration
96-hr
96-hr
35-day
Endpoint
LC50
LC50
LOEC
NOEC
MATC
Concentration
(mg/L)
1.1
1.1
0.31
0.16
0.22
Chemical
Analysis
Measured
Flow-through
Nominal
Static
Flow-through
Measured
Effects
Mortality
Mortality
Early Life Stage
Survival,
Hatching
Reference
ACC (2003)
Chow etal. (2013)
SLS (1989c)
Note: The shaded rows indicate the principal study used for assessing acute / chronic risks to aquatic organisms.
a Test guideline/type not reported
LOEC = Lowest Observed Effect Concentration
NOEC = No Observed Effect Concentration
MATC = Maximum Acceptable Toxicant Concentration
OC = Organic Carbon
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F-2 Toxicity to Terrestrial Organisms
F-2-1 Terrestrial Plant Toxicity
The effect of tetrabromobisphenol-A (purity: 99%) on 21-day emergence and growth of six species of
plants (ACC-BFRIP, 2002) was determined using OECD Guideline 208 (proposed version), US EPA OPPTS
850.4100 (1996) and US EPA OPPTS 850.4225 (1996).
The following six plant species were tested: Monocots - corn (Zea mays), onion (Allium cepa), ryegrass
(Lolium perenne); Dicots - cucumber (Cucumis sativa), soybean (Glycine max) and tomato (Lycopersicon
esculentum). Overall, treatment-related effects on seedling growth were seen in five out of the six
species tested (soybean was the exception). In contrast, no treatment-related effects were seen on
seedling emergence or condition of seedling in any of the species tested. The MATCs for corn,
cucumber, onion, ryegrass and tomato are 518, 32, 518, 127 and 518 mg/kg dry soil, respectively (ACC-
BFRIP, 2002).
Terrestrial plants were not considered for the proposed TBBPA risk assessment. For perspective, TBBPA
is two to three orders of magnitude less toxic to terrestrial plants than to soil-dwelling organisms.
F-2-2 Soil Invertebrate Toxicity
EPA considered three studies evaluating potential toxicity to soil-dwelling organisms to be of sufficient
quality to support hazard characterization of TBBPA. These studies are summarized in Table_Apx F-2.
In 2003, earthworms (Eisenia fetida) were exposed to TBBPA in a 56-day reproductive study (ACC-
BFRIP, 2003). The experimentally-derived and analytically-measured 56-day NOEC, LOEC and MATC
values were 2.1, 4.5 and 3.1 mg/kg soil (dry weight), respectively. In 2005, results from a 56-day
reproductive study with soil-dwelling worms (E. fetida) indicated that adverse effects from TBBPA were
observed at nominal concentrations as low as 0.63 mg/kg (LOEC) (ACC-BFRIP, 2005a). The reported
nominal NOEC is 0.31 mg/kg dry weight soil and the MATC was determined to be 0.44 mg/kg dry
weight soil. Furthermore, an additional earthworm study was conducted in 2006 using reproduction as
the endpoint of concern (Sverdrup et al., 2006). Sverdrup et al. (2006) exposed Enchytraeus crypticus
to TBBPA for 21 days. Based on reproduction, the experimentally-derived and analytically-measured
21-day NOEC, LOEC and MATC values were 3, 10 and 5.5 mg/kg soil (dry weight), respectively.
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Table_Apx F-2: Soil Invertebrate Toxicity Data for TBBPA (mg/kg)
Test Organism
Test Guideline/
Study Type
Duration
(day)
Endpoint
Value
(mg/kg)
Test
Analysis
Effect
Reference
Terrestrial Organisms
Earthworm
(Enchytraeus
crypticus)
Earthworm
(Eiseniafetida)
Earthworm
(Eiseniafetida)
ISO, 2002
USEPAOPPTS 850.6200,
OECDTG207
USEPAOPPTS 850.6200,
OECDTG 207/222
21-day
56-day
56-day
LOEC
NOEC
MATC
LOEC
NOEC
MATC
ECso
LOEC
NOEC
MATC
10
3
5.5
4.5
2.1
3.1
0.91
0.63
0.31
0.44
Measured
Measured
Nominal
Reproduction
Reproduction
Reproduction
Sverdrup et al.
(2006)
ACC-BFRIP
(2003)
ACC-BFRIP,
2005a
Note: The shaded row indicates the principal study used for assessing acute risks to aquatic organisms.
LOEC = Lowest Observed Effect Concentration
NOEC = No Observed Effect Concentration
MATC = Maximum Acceptable Toxicant Concentration
The most sensitive species is expected to be protective of a wider variety of species that are not
specifically represented by the available experimental data. The E. fetida 56-day MATC of 0.44 mg/kg,
based on reproduction, was the lowest endpoint, and it was divided by an uncertainty factor (UF) of 10
for chronic effects, as per established EPA/OPPT methods (EPA, 2012d, 2013b), to give a chronic COC
of 0.044 mg/kg or 44 u.g/kg.
F-2-3 Avian Toxicity
Under EPA's High Production Volume (HPV) Chemical Challenge program, data on birds have been
submitted for TBBPA (ACC, 2006) and are described below.
Several experiments have been conducted on distribution and effects of TBBPA in birds. 14C-TBBPA
was injected into quail eggs (1.9 u.g/g egg) on day 3 of incubation, and uptake and distribution of 14C-
TBBPA was studied in 6- and 9- day-old quail embryos. TBBPA was also administered to adult females
(single oral or intravenous doses, 250 jag/bird) to investigate its distribution. In addition, the potential
for certain reproductive and endocrine effects were evaluated in adult birds after embryonic exposure
(15 u.g/g egg). The embryonic uptake of TBBPA was low (< 1% of the radiolabel) after yolk injection on
day 3 and was distributed in the yolk, although metabolism was detected based on labeling in the liver,
bile and allantoic fluid. Thus, TBBPA's transfer to the embryo from the yolk was low with rapid
metabolism and excretion (Halldin et al., 2001).
In laying quail, TBBPA was rapidly eliminated via bile and excreted in feces, and transfer to egg yolks
was low. This effect was seen after oral and intravenous administration. In ovo exposure to TBBPA (15
u.g/g egg) did not cause estrogen-like effects in the adult quail. Egg-laying was not affected in female
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birds, and no effect in male quail on sexual behavior, testis weight or plasma testosterone was
detected (Halldin et al., 2001).
A previous study by the same group of investigators after 45 u.g TBBPA /g was injected in quail eggs
also did not find estrogenic-like effects (Berg et al., 2001).
Given the lack of effects observed in these studies and method of exposure (egg injection) for some
studies, EPA/OPPT doesn't plan to evaluate these effects in the proposed risk assessment.
F-3 Toxicity to Amphibians
Data on amphibians have also been submitted to EPA for TBBPA (ACC, 2006) and have been evaluated
by EC/HC (2013).
In the frog embryo teratogenesis assay: Xenopus (FETAX) bioassay, TBBPA was evaluated for potential
hormonal activity in the tadpole (Xenopus) embryo during the first 96 hours of development. Mortality,
malformation rate and growth inhibition/acceleration were evaluated by measuring changes in embryo
length and the presence of features indicative of earlier/later life stages. At 0.1, 1, 10 100 or 500 u.g/L,
TBBPA had no effect on Xenopus development when using either standard or minimal levels of sodium
and potassium (Garber et al., 2001).
TBBPA was also evaluated in the tadpole (Xenopus) tail regression assay to determine possible effects
on thyroid hormones. At developmental stage 5818, TBBPA was microinjected into tadpoles at doses up
to 60 u.g/tadpole. Although positive controls showed delayed tail resorption, TBBPA showed no effects
on tail resorption (Balch and Metcalfe, 2001).
TBBPA was found to decrease functioning of the thyroid hormone, triiodothyronine (T3), which is
critical to the triggering and control of metamorphosis in amphibians (Brown etal., 1996; Hanadaetal.,
2003; Kashiwagietal., 1999). Veldhoen et al. (2006) studied pre-metamorphic tadpoles of the Pacific tree
frog (Pseudacris regilla) and found that normal thyroid hormone-mediated gene expression profiles
were significantly altered at both TBBPA concentrations (5.4 jag/Land 54 u.g/L) evaluated. The results
show changes in endocrine-regulated gene expression at a sensitive life stage of the frog can occur
within hours of exposure to low concentrations (EC/HC, 2013).
The possible adverse effects of tetrabromobisphenol A exposure on the endocrine system in
amphibians have shown mixed results. Furthermore, the effect of changes in gene expression is
not clear. For these reasons, EPA/OPPT has not considered these results further for inclusion
in a risk assessment of TBBPA.
18 Normal growth at this stage would be the following: hind limbs emerged and forelimbs formed but not emerged.
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F-4 Summary of Environmental Hazard
Ecotoxicity studies for TBBPA have been conducted in aquatic plants, aquatic invertebrates (water
column and sediment), fish, terrestrial plant species, soil invertebrates and birds. Four COCs were
derived for the assessment, as summarized in Table_Apx F-3. The first COC in the table was
determined by dividing the acute effect level by an uncertainty factor of 5; the three chronic COCs
were calculated using a chronic uncertainty factor of 10.
Table_Apx F-3: Ecotoxicity Concentrations of Concern
Environmental Toxicity
Acute toxicity, aquatic water
column organisms
Chronic toxicity, aquatic
water column organisms
Chronic toxicity,
aquatic/sediment-dwelling
organisms
Chronic toxicity, terrestrial
organisms
Concentration of Concern
(COC)
0.02 mg/L
0.002 mg/L
11.7 mg/kg
0.044 mg/kg
Species and Effect
Eastern oyster;
shell deposition
Blue mussel;
Growth rate,
shell length
Blackworm;
reproduction
Earthworm;
reproduction
Reference
SLS (1989a)
ACC (2005)
Krueger(2002a)
ACC-BFRIP(2005)
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Appendix G Human Health Hazard Study Summaries
The toxicological effects of TBBPA have been extensively reviewed and evaluated by the European
Union (EC, 2006) and Health Canada (EC/HC, 2013). EPA/OPPT's Design for the Environment and the
National Toxicology Program (EPA, 2014e; NTP, 2014a) have also provided toxicity reviews. Therefore,
most of the sections below that describe available toxicokinetics and toxicodynamics studies as well as
mode of action data rely heavily on previous assessments. The reader is referred to these previous
reviews and original articles for more detailed information. A more recent study, the NTP (2014a)
cancer bioassay, is described in more detail.
Most studies using TBBPA have been rodent studies, primarily conducted via the oral route. In a recent
NTP cancer bioassay, TBBPA was associated with more than one tumor type in rodents. Also, some
developmental/newborn rodent studies resulted in effects. However, many of the subchronic,
reproductive and developmental toxicity studies conducted in rodents found no effects at the highest
doses tested.
G-l Epidemiology
In a cross-sectional study on 515 high school students in Belgium, TBBPA measured in serum was not
consistently associated with changes in neurobehavioral function in these adolescents, as assessed by
several types of tests (Kicinski et al., 2012).
G-2 Toxicokinetics
There are no in vivo toxicokinetics data via the dermal route for TBBPA. However, in vitro data show
that TBBPA applied as 2 mg/cm2 to human skin for 24 hours results in limited absorption (< 1%) (ECHA,
2013). This information, combined with TBBPA's low water solubility, high molecular weight and high
log Kow suggest that the compound will have limited absorption through the skin (EC, 2006).
Data are also lacking on TBBPA's toxicokinetics after inhalation. TBBPA has a median particle size of
approximately 31.81 to 52.20 u.m. Approximately 4% of the particles are <15 u.m in diameter (EC,
2006). The particle size distribution indicates that only a small amount of TBBPA is expected to be
respirable (< 10 u.m) and an even smaller amount of this respirable fraction is expected to deposit in
the alveolar region of the lungs after inhalation. Another amount will deposit in the nasopharyngeal
region and be either exhaled or swallowed/absorbed through the Gl tract. The EU assumed that
approximately 75% of inhaled particles will be absorbed; 70% of this was assumed to be from
swallowing/GI tract absorption and 5% from absorption through the lung (EC, 2006).
Several oral toxicokinetics studies have been conducted in rodents. An oral study of Sprague Dawley
rats shows that considerable amounts (e.g., > 70%) of TBBPA can be absorbed by the gastrointestinal
tract (Hakk et al., 2000). TBBPA can undergo Phase II metabolism directly, primarily conjugation via
glucuronidation and/or sulfation. The conjugated forms are then excreted primarily via the feces after
oral exposure (Hakk et al., 2000; Kuester et al., 2007; Schauer et al., 2006).
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Approximately 95% of TBBPA or its metabolites were excreted in the feces within 72 hours after a
single oral administration and only about 1% was excreted in the urine (Hakk et al., 2000). There was
limited retention of TBBPA or metabolites in the blood and tissues (liver: 0.4%, muscle: 0.12%, skin:
0.12%, fat: 0.7%, blood: 0.01%). Rapid elimination and low tissue retention of TBBPA were also
reported in Fischer 344 rats following oral or intravenous routes of exposure. Rates of elimination and
tissue retention were comparable for single doses or for repeated daily doses of 20 mg/kg-bw for 5 or
10 consecutive days (Kuester et al., 2007). One study in pregnant rats found limited transfer of TBBPA
or its metabolites to the fetus; tissues in the pregnant rats and fetuses contained about 1.2% and
0.34%, respectively, of radioactivity 48 hours after exposure to 14C ring-labeled TBBPA (Meerts et al.,
1999).
Similar to results of rat studies, TBBPA-glucuronide and TBBPA-sulfate were identified as the major
metabolites in humans after oral administration of TBBPA (Schauer et al., 2006; Zalko et al., 2006). The
half-life of TBBPA in humans has been estimated to be about 2 days (Hagmar et al., 2000; Sjodin et al.,
2003).
G-3 Acute Toxicity
Acute oral, dermal and inhalation studies have been performed with TBBPA. These studies show that
TBBPA has low acute toxicity by all routes of exposure. The oral LDsos are > 50,000 mg/kg-bw in rats
and 3200 mg/kg-bw in mice; the dermal LDso is > 10,000 mg/kg-bw in rabbits; and inhalation LCsos are
> 10.92 mg/L in rats and > 50 mg/L in mice (GLCC, 1967; Gustafsson and Wallen, 1988; HTRI, 1966;
IBRI, 1967; VCC, 1978a).
Nakajima et al. (2009) evaluated neurobehavioral changes in an open-field test 3 hours after gavage
exposure using 0.1, 5 or 250 mg/kg-bw/day TBBPA. Some behavioral effects were seen at 0.1 and 5.0
but not at 250 mg/kg-bw/day. TBBPA was found in the striatum at the two lowest doses and showed
non-specific accumulation in the brain at 250 mg/kg-bw/day. Given the lack of a dose response, it is
difficult to make a conclusion regarding acute neurobehavioral effects from this study.
G-4 Repeated-Dose Toxicity
The chronic and subchronic toxicity of TBBPA was investigated in an inhalation study, a dermal study
and several oral repeated-dose studies. The inhalation and dermal studies are discussed. Oral studies
of 90 days or longer are also summarized.
G-4-1 Inhalation
In a 14-day inhalation study in rats (IRDC, 1975), the EU concluded that no adverse systemic effects
were reported at 2, 6 and 18 mg/L, with the exception of signs of mechanical irritation in all treatment
groups due to the high dust levels (EC, 2006).
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G-4-2 Dermal
In a 21-day dermal study, TBBPA was administered to rabbits at doses up to 2,500 mg/kg-bw/day for 6
hours/day, 5 days/week. No toxicologically significant effects were identified (IRDC, 1979).
G-4-3 Oral - Gavage
In a 90-day oral gavage study, rats were administered TBBPA at doses of 0, 100, 300 or 1,000 mg/kg-
bw/day (MPI_Research, 2002a). No neurobehavioral effects were observed during the weekly
functional observational battery evaluations. Slight changes in hematological evaluations and clinical
chemistry were reported; however, the EU concluded that these effects were not toxicologically
significant. Statistically significant decreases in serum T4 were reported in males and females, but no
accompanying change in serum Ts, thyroid stimulating hormone (TSH) or histopathology of the liver,
thyroid, parathyroid or pituitary was reported. The EU again concluded that the decreases in serum T4
were not adverse. Absolute spleen weight was decreased in males in the top two dose groups; no
histopathological findings were noted. An increase in relative epididymis weight was reported in the
middle dose group; however, no changes in relative epididymis weight or histopathology were
identified in the high dose group. Again, the EU concluded that these findings were of no toxicological
significance (EC, 2006).
In a recent subchronic toxicity study, groups of 10 male and 10 female F344/NTac rats and B6C3F1
mice were administered 0, 10, 50, 100, 500 or 1,000 mg/kg-bw/day TBBPA in corn oil by gavage, 5 days
per week for up to 14 weeks (NTP, 2014a). In the rats, dose-related decreases in serum T4 levels
occurred on day 4 and at week 14 in 500 and 1,000 mg/kg-bw/day males and females. These effects
occurred less consistently at 100 mg/kg-bw/day. Significant increases occurred in liver weights of 500
and 1,000 mg/kg-bw/day rats (males and females) and significant decreases occurred in spleen weights
of 500 and 1,000 mg/kg-bw/day males. No treatment-related histopathologic lesions were observed.
In the mice, liver weights of 500 mg/kg-bw/day males and 1,000 mg/kg-bw/day males and females
were significantly greater than those of the vehicle controls. Kidney weights were significantly
decreased and spleen weights were significantly increased in 1,000 mg/kg-bw/day males. At 500 and
1000 mg/kg-bw/day, males exhibited increased incidences of renal tubule cytoplasmic alterations. No
additional treatment-related histopathologic lesions were observed in mice in this 14-week study (NTP,
2014a).
Administration of TBBPA in corn oil to Wistar Han rats at 0, 250, 500 or 1,000 mg/kg-bw/day 5 days per
week by gavage for up to 105 weeks resulted in decreased body weight (by at least 10% lower than
vehicle controls after week 25) at 500 and 1,000 mg/kg-bw/day. At three months, thymus weights in
the 1,000 mg/kg-bw/day dose group were significantly lower and liver weights were higher than
vehicle controls. Females at all doses exhibited increased incidences of nonneoplastic lesions of the
uterus (NTP, 2014a).
B6C3F1 mice (50/sex/dose) were also administered TBBPA via gavage in corn oil at 0, 250, 500 or 1,000
mg/kg-bw/day for 5 days/week for up to 105 weeks. Mice at 1,000 showed decreased body weights
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(by more than 10% after week 25) and decreased survival compared with vehicle controls. In male
mice, nonneoplastic lesions were seen in the liver and kidney at 250 and 500 mg/kg-bw/day. Male
mice at 500 mg/kg-day and female mice at 250 and 500 mg/kg-bw/day exhibited forestomach lesions
(NTP, 2014a).
G-4-4 Oral - Dietary
Rats were fed a diet for 3 months that included approximately 0, 0.3, 3, 30 or 100 mg/kg-bw/day
TBBPA. No changes were observed in clinical signs, body weights, hematology, clinical chemistry,
urinalysis, organ weights, gross or microscopic pathology (Dow_Chemical, 1975).
TBBPA was given to B6C3F1 mice at 0, 500, 4900, 15,600 or 50,000 ppm (approximately 0, 71, 700,
2,200 or 7,100 mg/kg-bw/day) for 3 months. All mice at 7,100 mg/kg-bw/day died but no mice died at
lower doses. Body weight gains were decreased at 2,200 and 7,100 mg/kg-bw/day and red blood cells,
hemoglobin, hematocrit, serum triglycerides and total serum proteins were decreased at 2,200 mg/kg-
bw/day. Increased spleen weights were also observed (Tobe et al., 1986).
G-5 Reproductive and Developmental Toxicity
G-5-1 Multigenerational Studies
No effects on reproduction, fertility or developmental toxicity including neurobehavioral abnormalities
(e.g., motor activity, learning and memory, auditory response) were observed in an OECD compliant
two-generation study in Sprague-Dawley rats using gavage doses of 0,10, 100 or 1,000 mg/kg-bw/day
(MPI_Research, 2002b, 2003). In this study, there were significant decreases in serum thyroxine (14)
levels in Fo and Fi offspring at 100 and 1,000 mg/kg-bw/day. Mean serum Ts levels were also
significantly lower in Fo males at the highest dose (1,000 mg/kg-bw/day), but no changes were found in
the Fo females or in the male or female Fi offspring. There were also no effects on TSH levels or
microscopic changes in the liver or pituitary gland; the thyroid was not examined histopathologically.
There was no dose-response relationship in T4 levels; following the 30-day recovery period, the levels
were similar to controls. The European Union concluded that the thyroxine effects were not
toxicologically significant (EC, 2006).
Van der Ven et al. (2008) reported the results of a one-generation reproduction dietary study in Wistar
rats using TBBPA doses of ~ 0, 3, 10, 30, 100, 300,1,000 or 3,000 mg/kg-bw/day. In this study,
exposure lasted 11 weeks (males) and 2 weeks (females) prior to mating and during mating (both
sexes). For females, dosing continued throughout gestation and lactation. After weaning, dosing of
offspring were 10 weeks old. Dams exhibited decreased body weight at the highest dose along with
reduced food consumption. However, benchmark doses calculated at the lower 95% confidence limit
for an associated 10% decrease in body weight were close to the highest dose and sometimes higher
than 3000 mg/kg-bw/day. There were no effects on the reproduction parameters examined, including
sperm count or morphology.
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In offspring, decreased plasma T4 levels were associated with lower bounds on benchmark doses
(BMDLios) of 30.8 and 16.1 mg/kg-bw/day in males and females, respectively. Modeling of the
increased Ts levels in female offspring resulted in a BMDLio of 2.3 mg/kg-bw/day; this effect was not
seen in males. The authors also report BMDLios of 0.5 and 0.6 mg/kg-bw/day, respectively, for
increased testicular and pituitary gland weights in male offspring. The testes, pituitary gland and
thyroid gland did not exhibit histopathological changes to accompany the increased organ weights or
hormone changes. Other effects in offspring (decreased anogenital distance in females at day 7 but not
day 4 or 21; number of days to vaginal opening) were observed. However, BMDLs for these effects
were calculated as 2736 and 2745 mg/kg-bw/day, respectively (Van der Ven et al., 2008). Review of
data by dose level as presented in this publication didn't always reveal clear dose-response
relationships.
The possible neurobehavioral effects in offspring from the above one-generation reproduction study
were investigated from PND 50 to 140 after being exposed in utero and after direct dosing of TBBPA.
The authors examined auditory responses by measuring brainstem auditory evoked potentials (BAEPs),
which are electrophysiologic responses elicited by auditory stimuli and recorded from the scalp or
brain surface as waveform with a series of positive and negative peaks (Lilienthal et al., 2008).
The authors reported that BAEP thresholds and wave IV latency were increased in exposed female
offspring in the low sound frequency range. In the males, absolute latency of wave IV and interpeak
latencies II-IV were also increased at low frequencies of sound. The authors reported a BMDLio of 8
mg/kg-bw/day for wave IV latency. BMDLs for increased BAEP thresholds in females ranged from 1 to
40 mg/kg-bw depending on the sound frequency at which they were measured (Lilienthal et al., 2008).
Both of the above studies have been criticized for various reasons. Banasik et al. (2009) expressed
concerns about Van der Ven et al. (2008) regarding the use of modeling software, methodology and
conduct of the study. Also, the effects identified by Van der Ven et al. (2008) were not considered
critical endpoints by Health Canada (EC/HC, 2013). Some of the methods and statistical analyses of the
findings presented by Lilienthal et al. (2008) were also called into question by Strain et al. (2009).
Other limitations or deficiencies of Lilienthal et al. (2008) were identified by comparing the study with
EPA's Office of Chemical Safety and Pollution Prevention (OCSPP) Test Guidelines on Neurophysiology:
Sensory Evoked Potentials (870.6855, August 1998). According to these EPA test guidelines, a
pigmented strain of rat is the preferred animal species to be tested because albino strains of animals
have known abnormalities of the visual and auditory systems. Furthermore, at least 10 nulliparous and
nonpregnant rats per group should be used, and positive control groups exhibiting functional changes
in the sensory systems to be tested are recommended. Instead, the study of Lilienthal et al. (2008)
used groups of 5-6 pregnant Wistar rats (which are albino rats) and did not include positive controls.
In a separate review of Lilienthal et al. (2008) by an EPA neurotoxicologist, it was noted that although
the criticisms by Strain et al. (2009) have some merit, they are not sufficient to completely dismiss the
findings (Herr, 2013). EPA/OPPT believes that this study suggests the potential for auditory effects but
will not use this study in a quantitative risk assessment given some concerns (one being the use of
albino rats) and the difficulty determining what exposures may lead to the effects (e.g., whether
exposure by the dams/parents or direct exposure to the offspring experiencing the effects is most
relevant).
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Zatecka et al. (2013) conducted a 2-generation drinking water study in GDI outbred mice. In the
parental generation, TBBPA was administered only to pregnant dams during gestation (at 0 or 0.035
mg/kg-bw/day) and not to fathers. Fi offspring were then dosed during pre-pubescence, pubescence
and up to adulthood and evaluated at 70 days of age. In the Fi offspring, there was an increased
incidence of apoptotic cells in testes and increased expression of genes encoding proteins important
during spermatogenesis. The Fi generation was then bred in a cross-over fashion using four patterns
(both parents exposed; neither exposed; mother exposed; father exposed). The F2 offspring were then
evaluated at 70 days of age; it is not clear from the study description whether the F2 rats had access to
TBBPA in the drinking water or from dam's milk or both. In the F2 offspring with both parents exposed
to TBBPA, testicular weights were reduced (p < 0.01), prostate weights were increased (p < 0.05) and
seminal vesicle weights were increased (p < 0.01). The F2 generation with only fathers exposed showed
increased epididymis weights (p < 0.01). No visible abnormalities or pathological changes in
seminiferous tubule morphology were seen. The significance in these effects is unclear because only
one dose was used and histopathological, sperm and other reproductive effects were not observed.
G-5-2 Prenatal and Postnatal Developmental Toxicity Studies
There were no significant dose-related effects on T4, Ts orTSH in offspring at PND 20 or postnatal week
11 after pregnant Sprague-Dawley rats were fed TBBPA at dietary levels of 0, 100, 1,000 or 10,000 ppm
(~ 17,149 or 1472 mg/kg-bw/day) from gestation day (GD) 10 to PND 20. However, offspring exhibited
slight decreases in Ts on PND 20 that were not dose-related. TBBPA did not alter brain development or
other reproductive parameters. The study did not state whether offspring had direct access to TBBPA
in the food (Saegusa et al., 2009).
In a pilot range-finding study (VCC, 1978b), no developmental effects were observed in offspring of
dams administered TBBPA up to 10,000 mg/kg-bw/day from GD 6-15. In a standard developmental
toxicity study (Noda et al., 1985), pregnant Wistar rats were administered TBBPA via gavage from GD 0
to 20 at doses up to 2,500 mg/kg-bw/day. No effects were seen in dams or in offspring evaluated up to
PND 21. In another developmental toxicity study (MPI_Research, 2001), pregnant rats were
administered TBBPA via gavage up to 1,000 mg/kg-bw/day from GD 0 to 19 and again, no effects were
observed in dams or offspring.
No differences in various neurobehavioral measures in a study of NMRI mice administered 0.75 and
11.5 mg TBBPA/kg-bw orally on PND 10 (Eriksson et al., 2001; Eriksson et al., 1998).
Viberg and Eriksson (2011) report biochemical changes related to cholinergic effects in neonatal NMRI
mice treated with 11.5 mg TBBPA/kg-day. Radioactivity from TBBPA dosing was highest (at 3.7%) in the
brain at 3 hours, decreased to 0.9% at 24 hours and was 0.3% by 7 days after dosing. It is not entirely
clear whether the study controlled for possible litter effects. Although some evaluations were done
with equal numbers of animals per litter, methods of litter culling were less clear for other assessments
(Viberg and Eriksson, 2011).
In a study in which TBBPA was administered at 0%, 0.01%, 0.1% or 1% in the diet to pregnant ICR mice
(6/dose) from the first day of gestation to weaning at postnatal day 27, no effects on average litter
size, litter weight, total number of offspring, average male or female offspring weights or dam weights
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were reported (Tada et al., 2006). The offspring drank the dam's milk while the dams were being
treated. Also, offspring had access to dam's food from PND 22 to 27; it is not known how much was
ingested. Changes in total cholesterol, triglycerides and organ weights (liver, brain, spleen) were seen
primarily at the highest dose (p < 0.05) in the dams. Female offspring exhibited an increased incidence
of renal tubule atrophy at the highest dose. Based on enlargement of hepatocytes and very slight focal
necrosis of hepatocytes in female offspring, Health Canada considered the LOAEL to be 0.1% (~140.5
mg/kg-bw/day during gestation), with a NOAEL of 0.01% (~15.7 mg/kg-bw/day during gestation).19
Littermates were used as independent variables for the experimental and statistical analyses. Thus, the
tendency of littermates to respond more similarly to one another than non-litter mates was not fully
taken into account. However, the authors chose the same number (4 offspring of each sex) randomly
from each litter to partially account for such effects (Tada et al., 2006).
In a dietary developmental toxicity study, Saegusa et al. (2012) administered TBBPA to Sprague Dawley
rats from GD 10 through PND 12. The highest concentration (10,000 ppm or ~ 800 mg/kg-bw/day)
resulted in increased interneurons in the dentate hilus-expressing reelin; this effect suggests
alterations in neuronal migration. It is not clear from this study whether offspring had direct access to
TBBPA in the diet of the dams after birth.
In a study of newborn rats directly dosed by gavage from days 4 to 21 after birth, an effect on the
kidneys (polycystic lesions associated with the dilation of tubules) was noted at 200 and 600 mg/kg-
bw/day but not at 40 mg/kg-bw/day. Effects at 600 mg/kg-bw/day were considered moderate in
females and severe in males whereas the lesions seen in the two males that exhibited effects at 200
mg/kg-bw/day were of slight severity. At 85 days of age, nephrotoxic lesions were still seen at 200 and
600 mg/kg-bw/day. There were no neurobehavioral effects as assessed by testing reflexes on PND 21
(Fukuda et al., 2004). Based on the slight effects in the kidney at 200 mg/k-bw/day, the LOAEL is
considered to be 200 and the NOAEL 40 mg/kg-bw/day.
In comparison, no similar effect was found in 5-week old rats dosed via gavage with 2,000 or 6,000
mg/kg-bw/day for 18 days (Fukuda et al., 2004). The EU suggested that the kidney effects observed in
the newborn rats are likely due to the immature metabolic capability and/or immature kidneys (EC,
2006).
G-6 Irritation and Sensitization
The European Union and Canada concluded that TBBPA is not a skin, eye or respiratory irritant and is
not considered to be a skin or respiratory sensitizer in animals or humans (EC, 2006; EC/HC, 2013;
GLCC, 1967; Gustafsson and Wallen, 1988; HTRI, 1966; IBRI, 1967; VCC, 1978a).
G-7 Genotoxicity and Carcinogenicity
The following sections present the finding of the NTP (2014a) cancer bioassay and discuss conclusions
from the Office of Pesticide Program's (OPP's) Cancer Assessment Review Committee (CARC) about
19 The LOAEL and NOAEL during lactation are associated with doses of 379.9 and 42.1 mg/kg-bw/day, respectively.
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TBBPA-related tumors and human relevance. Mode of action is also discussed along with EPA/OPPT's
overall classification regarding human relevance.
G-7-1 Genptpxicity
Both in vitro and in vivo assays discussed in available reviews have found that TBBPA is not genotoxic.
These studies include multiple bacterial reverse mutation assays with and without metabolic
activation, an in vitro test for intragenic recombination in mammalian cells, an in vitro mammalian
chromosomal aberration test using human peripheral blood lymphocytes (with and without metabolic
activation) and a mouse micronucleus study(EC, 2006; NTP, 2014a).
G-7-2 NTP Carcinpgenesis Bipassays
NTP recently published a report of the results of carcinogenesis studies of TBBPA in rats and mice (NTP,
2014a). Groups of Wistar Han rats and B6C3Fi mice were administered 0, 250, 500 or 1,000 mg/kg-
bw/day TBBPA in corn oil by gavage, 5 days per week for up to 105 weeks. At 0 and 1,000 mg/kg-
bw/day, 60 rats/sex/dose were used and at 250 and 500 mg/kg-bw/day, 50 rats/sex/dose were used.
Fifty mice/sex/dose were used for all doses.
G-7-2-1 Findings in Rats
Males
In males, mean body weights were at least 10% lower than controls after week 25. There were no
clinical findings and survival did not differ between dosed groups and controls. At three months, there
were no treatment-related lesions. After two years, the incidences of interstitial cell adenoma of the
testis were slightly increased at 500 (1/50) and 1,000 mg/kg-bw/day (3/50) compared to controls
(0/50) (p < 0.05 for the trend test). Pairwise comparisons were not statistically significant.
Females
Body weights were similar in all groups, including controls. There were no clinical findings and survival
was similar among all groups. No treatment-related lesions were seen at the three-month evaluation.
For the uterus only (not other tissues), NTP first evaluated the tissues transversely and later evaluated
them longitudinally. According to the original transverse evaluation, there were increases in both
neoplastic and non-neoplastic lesions of epithelial origin. Cystic endometrial hyperplasia of the
endometrium was increased at 1,000 mg/kg-bw/day (p < 0.05 by the Poly-3 test). NTP found positive
dose-response trends in the incidences of 1) adenoma (p = 0.001) and 2) adenocarcinoma (epithelial
origin) (p < 0.05; Poly-3 test). When tumors were combined, the incidences of adenoma,
adenocarcinoma or malignant mixed Mullerian tumors of the uterus showed a positive trend (p < 0.05
by the Poly-3 test). By pairwise comparison, the incidences were significantly increased in the 500 and
1,000 mg/kg-bw/day groups (p < 0.05 and < 0.01 by the Poly-3 test).The combined incidences of these
three tumor types in control, low-, mid- and high-dose groups were: 3/50, 7/50, 11/50 and 13/50,
respectively.
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More nonneoplastic and neoplastic lesions of epithelial origin were identified after longitudinal
evaluation of residual uterine tissue. When the original and residual evaluations were combined,
atypical endometrial hyperplasia was significantly higher than controls at all doses (p < 0.01; Poly-3
test). However, cystic endometrial hyperplasias were no longer significantly higher than controls. There
were positive trends (Poly-3 test) in the incidences of 1) adenocarcinoma alone (p < 0.01) and 2) the
combined incidences of adenoma, adenocarcinoma or malignant mixed Mullerian tumor (p < 0.001) for
the combination of original and residual sections. For both adenocarcinomas alone and for the
combined tumor types, pairwise comparisons showed that incidences were increased in the 500 and
1,000 mg/kg-bw/day groups (p <0.01; Poly-3 test). The combined incidences of the three tumor types
in the control, low-, mid- and high-dose groups were: 6/50, 11/50, 16/50 and 19/50, respectively.
In the ovaries, incidences of rete ovarii cysts were significantly greater at 500 and 1,000 mg/kg-bw/day
compared with controls (p < 0.05; Poly-3 test).
G:7:2-2 Findings in Mice
Both Sexes - Decreased Survival
Survival of mice at 1,000 mg/kg-bw/day (577 days for males; 413 days for females) was lower than
controls (687 days for males; 711 days for females) (p < 0.001). Decreased survival was seen as early as
6 months. Although (NTP, 2014a) reports numbers of 1,000 mg/kg-bw/day mice that had neoplasms,
the incidence is lower than the 500 mg/kg-bw/day group and the decreased survival at 1,000 mg/kg-
bw/day was not attributed to the tumors. Instead, decreased survival was associated with decreased
body weight gain and may have been due to gastrointestinal toxicity. Because of the decreased
survival at the highest dose, NTP (2014a) evaluated tumors only for the 0, 250 and 500 mg/kg-bw/day
dose groups.
Males
Several non-neoplastic lesions were observed in the kidney and forestomach. Renal tubule cytoplasmic
alterations were observed at 250 and 500 mg/kg-bw/day (< 0.01; Poly-3 test) and increased in severity
by dose. In the forestomach, the 500 mg/kg-bw/day males exhibited ulcers, mononuclear cell
infiltration, inflammation and epithelium hyperplasia (p < 0.05 or < 0.01).
Both nonneoplastic and neoplastic lesions were observed in the liver. Compared with the control
group, the number of male mice with clear cell foci was increased at 500 mg/kg-bw/day; eosinophilic
foci were significantly increased at 250 and 500 mg/kg-bw/day (p < 0.01 for all). The incidence of
multiple hepatocellular adenoma was significantly increased at 500 mg/kg-bw/day compared with
controls (p < 0.05; Poly-3 test). In addition, the incidences of hepatoblastoma (11/50) and the
combined incidence of hepatocellular carcinoma or hepatoblastoma (24/50) in 250 mg/kg-bw/day
males (11/50) were significantly greater than those in the vehicle controls (2/50 or 12/50, respectively)
(p < 0.01; Poly-3 test). Differences were not statistically significant at 500 mg/kg-bw/day or by trend
tests.
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Other increased incidences of tumors were observed in the intestines or in all organs. The combined
incidences (0/50, 0/50 and 3/50 at 0, 250 and 500 mg/kg-bw/day) of adenoma or carcinoma of the
large intestine (cecum or colon) occurred with a significant positive trend (< 0.05 by the Poly-3 test).
The incidences of 1) hemangiosarcomas or 2) combined hemangiomas/hemangiosarcomas (3/50, 5/50
and 9/50 at 0, 250 and 500 mg/kg-bw/day) when summing tumors for all organs occurred with
significant positive trends (p < 0.05 by the Poly-3 test). By pairwise comparison, hemangiosarcomas
were higher than controls at 1,000 mg/kg-bw/day (p < 0.05; Poly-3 test).
Females
Females exhibited ulcers, mononuclear cell infiltration, inflammation and epithelium hyperplasia in the
forestomach at 250 and 500 mg/kg-bw/day (females). No increased incidences of tumors were found
in female mice.
G-7-2:3 NTP Conclusion
Under the condition of these studies, NTP concluded that there was clear evidence of carcinogenic
activity of TBBPA in female Wistar Han rats based on increased incidences of uterine epithelial tumors,
which primarily included uterine adenocarcinomas. There was some evidence of carcinogenic activity
of TBBPA in male B6C3F1/N mice based on the increased incidences of hepatoblastoma. NTP
concluded that there was equivocal evidence of carcinogenic activity of TBBPA in male Wistar Han rats
based on slightly increased incidences of testicular adenoma. The large intestine neoplasms and
hemangiosarcoma (all organs) are considered equivocal findings. There was no evidence of
carcinogenic activity of TBBPA in female B6C3F1/N mice (NTP, 2014a).
G-7-3 CARC Conclusions Regarding Weight of Evidence for
TBBPA-Related Tumors
The results of independent reviews of cancer bioassays/carcinogenicity, often of pesticides, are often
subsequently peer-reviewed by EPA's Cancer Assessment Review Committee (CARC), which reviews
data and recommends a cancer classification. This classification will then determine how the Agency
regulates pesticides or other reviewed compounds. The committee also recommends methods to
quantify human health risk (http://www.epa.gov/pesticides/health/cancerfs.htm). During the problem
formulation of TBBPA, EPA's OPPT asked the CARC to provide an independent review of the newly
available NTP study to help inform the cancer assessment.
The CARC conclusions are based on whether clear statistically significant dose-response trends and
pairwise statistical comparisons can be seen, whether precursor lesions were observed and whether
tumors are clearly within historical control ranges.
Similar to the NTP conclusions, the CARC considered uterine tumors in rats to be clearly related to
TBBPA treatment. However, in contrast to NTP's conclusions, the CARC made the following
conclusions:
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• TBBPA treatment is not associated with testicular tumors in male rats and intestinal tumors
(other than the hemangiosarcomas
• TBBPA is associated with hemangiomas and hemangiosarcomas in male mice
• TBBPA is not related to hepatoblastomas in male mice based on the lack of either statistical
significance or a dose trend. Despite statistical significance at the highest dose, the liver
adenomas in male mice are not of concern from TBBPA exposure because incidences were
within the historical control range and there were no precursor lesions.
The CARC's conclusions for each tumor type is further delineated in Supplemental File 4 (CARC, 2014).
The CARC concluded that there is not enough evidence regarding the mode of action (MOA) to meet
the International Programme on Chemical Safety (IPCS) 2007 Human Relevance Framework. This lack
of clear MOA and unknown human relevance resulted in CARC deferring to the default science position
of the 2005 EPA Cancer Guidelines (EPA, 2005) that tumors are relevant to humans. More information
on the CARC's overall weight of evidence specifically regarding relevance to humans is discussed in
Section 2.6.4.5G-7-5.
G-7-4 Mode of Action Considerations for Cancer
TBBPA's possible MOAs for induction of tumors are not clearly understood. Although not an evaluation
of all available data, some considerations as to TBBPA's possible cancer MOAs are discussed here, with
an emphasis on MOAs related to uterine tumors.
Based on the toxicokinetics data, it is possible that liver effects (i.e., enlargement of hepatocytes,
increased liver weight and slight focal necrosis of hepatocytes) observed in adult rats and/or mice after
exposures to high doses of TBBPA for extended periods could be due to saturated metabolic capability
and diminished elimination/excretion of the compound.
Negative genotoxicity studies suggest that a direct genotoxic MOA is not considered likely for
tumorigenesis. The CARC also concluded that there is no likely concern for mutagenicity for TBBPA
(CARC, 2014).
Uterine tumors can arise in response to endogenous estrogen overstimulation in aged rats, which can
be exacerbated by administration of exogenous chemicals through direct and indirect pathways (Alison
et al., 1994; Lax, 2004). After binding directly to estrogen receptors (ERs) in a cell, endogenous
estrogen and estrogen agonists can activate hormone-responsive genes that promote DNA synthesis
and cell proliferation. Therefore, estrogen and estrogen agonists can act as tumor promoters by
inducing proliferation of cells with pre-existing mutations and eventually lead to tumor formation.
Estrogen (as a result of excessive exposure) has been recognized as a known human carcinogen (IARC
group 1 carcinogen) (NTP, 2014b).
The ER binding activity of TBBPA has been investigated in a number of in vitro screening assays. Review
of the overall weight-of-evidence from in vitro assays has indicated that there is no significant
estrogenic potential for TBBPA. A recent study in mice has shown that TBBPA was negative for
estrogenic responses by both subcutaneous injection and oral routes of exposure up to 1,000 mg/kg-
Pagell9ofl35
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bw/day (Ohta et al., 2012). Because in vitro and in vivo studies showed that TBBPA has no significant
estrogenic potential, TBBPA is unlikely to operate directly by the ER-mediated pathway for the
induction of uterine tumors.
Conjugations (glucuronidation and sulfation) are the major biotransformation pathways for excretion
of TBBPA in rats, and these pathways are shared by estrogen (Raftogianis et al., 2000). Competition for
glucuronosyl-transferases and/or sulfotransferases by TBBPA could indirectly result in higher levels of
estrogen and increased formation of estrogen-derived reactive radicals following exposure to high
concentrations of TBBPA (NTP, 2014a).
In addition to competing with enzymes that metabolize estrogen, a recent crystallographic analysis of
TBBPA and a related brominated compound suggests that TBBPA, in addition to being a weak agonist,
may bind to and actually inhibit sulfotransferases (e.g., SULT1E1) that metabolize estrogen (Gosavi et
al., 2013) thereby causing possible buildup of estrogen.
Besides the promotional and indirect genotoxic effects, estradiol (and its interconvertible metabolite
estrone) can also exert genotoxic effects after being metabolized to catechols and then to reactive
quinones that can form DNA adducts and contribute to oxidative DNA damage by reactive oxygen
species (Bansal et al., 2009; Chen et al., 2008; Russo and Russo, 2004; Yager, 2014). If unrepaired,
these mutations may lead to tumor formation. There are data that suggest that both estrogen activity
and oxidative stress are required to induce cancer (Conova, 2003).
The Tp53 tumor suppressor gene is responsible for cell cycle checkpoint maintenance and genomic
stability, and loss of cell cycle checkpoint control due to Tp53 mutations can result in the development
of various tumor types in rodents and humans (Blagosklonny, 2000; Muller and Vousden, 2013). In the
NTP bioassays (NTP, 2014a), a statistically significant increase in the incidence of mutations of the Tp53
tumor suppressor gene was noted in uterine adenocarcinomas from TBBPA treated rats (60%)
compared to the incidences in mutations associated with spontaneous tumors from control rats (20%).
It is possible that increased incidence of Tp53 mutations may be caused by the reactive oxygen radicals
or metabolites produced after metabolism of the high levels of circulating estrogens due to
competitive binding/inhibition of high doses of TBBPA to sulfotransferases or glucuronosyl-
transferases. However, with the exception of multiple mutations in uterine tumors of two rats treated
with TBBPA, there was no difference between the mutation spectra of spontaneous tumors and those
from TBBPA-treated rats. Therefore, it may be more likely that increased estrogen levels due to
competitive inhibition of estrogen conjugations by high doses of TBBPA may cause uterine tumors by
promoting pre-existing Tp53 mutations in the uterus by turning on hormone-responsive genes that
promote DNA synthesis and cell proliferation.
Uterine/endometrial cancer is one of the most common cancers in women with over 54,000 new cases
estimated for 2015 (ACS, 2015).There are two types of uterine carcinoma with respect to histology,
MOA and molecular genetic pathways. Type I carcinoma (the most common type), is associated with
expression of ER, estrogen overstimulation, endometrial hyperplasia and Tp53 mutations in only about
10-20 % of the carcinoma. Type II carcinoma is unrelated to estrogen and frequent lack of estrogen
receptor activities. It is associated with atrophic endometrium, and Tp53 mutations (90%) are the most
frequent genetic alterations (Lax, 2004). In light of the high incidence of Tp53 mutations in the uterine
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adenocarcinomas, weak ER binding potential and the lack of increased levels of circulating estrogen in
TBBPA-treated rats, it is likely that TBBPA induced type II carcinoma in the rats (rather than type I
carcinoma). This notion is supported by the findings that there was no increase in endometrial
hyperplasia (which is associated with type I carcinoma) in all dosed groups of female rats in the NTP
bioassays when the original and residual tissues evaluations were combined; instead a new atypical
hyperplasia was identified (NTP, 2014a). The two types of uterine/endometrial carcinomas seen in the
rat study are similar to those observed in humans (e.g, Bansal et al., 2009) suggesting that the rat
model is relevant to humans.
Another possible MOA for TBBPA's association with uterine tumors deserves consideration. An in vitro
study investigating the effects of TBBPA on the uptake of neurotransmitters into isolated rat brain
synaptosomes showed a mixed concentration-dependent competitive/non-competitive mode of
inhibiting dopamine uptake (Mariussen and Fonnum, 2003). Uterine tumors can be induced by
dopamine receptor agonists in the rat through such an indirect estrogen pathway. Dopamine serves as
a key regulator of serum prolactin by activating dopamine receptors on the pituitary to inhibit the
secretion of prolactin. Chronic administration of dopamine receptor agonists to rats can result in
decreased serum prolactin levels after competitively binding to the dopamine receptors on the
pituitary, leading to estrogen dominance. Such dominance is due to increased estrogen synthesis after
luteolysis of the persistent corpora lutea and the formation of new follicles. This estrogen dominance
then leads to estrogen activity (i.e., expression of hormone-responsive genes that promote cell
proliferation) and oxidative stress, which may induce hyperplasia and tumors of the uterus. This
carcinogenic effect has not been demonstrated in other species including humans because prolactin is
the luteotrophic hormone in rodents but not in primates (Alison et al., 1994; Neumann, 1991).
Although the data above demonstrate TBBPA antagonism of dopamine uptake, it is unknown if TBBPA
can also act as a dopamine agonist.
NTP is now conducting additional studies that evaluate treatment-related molecular changes in the
uterus (NIEHS, 2015).
As noted in Section G-7-3, the CARC concluded that available mode of action data were not adequate
to meet the International Programme on Chemical Safety (IPCS) Human Relevance Framework (IPCS,
2007). The CARC did not discuss the specific information that was lacking to establish an MOA.
However, there are several important considerations identified by the Human Relevance Framework in
determining an MOA for any compound. The Framework specifies that the postulated key events
critical to the induction of tumors should be measured consistently. Concordance of dose-response
relationships between the key events and tumors is also needed, with consideration of Bradford Hill
criteria and whether differences in biological response (such as dose transitions) may occur at different
sections of the dose-response curve. Temporally, postulated key events for a mode of action should be
observed before the tumors are seen. There should be enough information to suggest some strength,
consistency and specificity when associating key events with tumor incidence. Biological plausibility is
important, as is the consideration of whether alternate modes of action occur (IPCS, 2007).
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G-7-5 Classification of Carcinogenic Potential for Humans
The weight of evidence for whether TBBPA is associated with tumors in rodents was presented in
2.6.4.5G-7-3. This section (G-7-5) specifically discusses conclusions regarding the relevance of these
data to humans.
The CARC classified TBBPA as likely to be carcinogenic to humans according to criteria described in
EPA's 2005 Cancer Guidelines. Some criteria that can influence a classification decision include the
presence of: 1) tumors in more than one species and sex, 2) more than one tumor type, 3) uncommon
tumors, 4) dose-response relationships, 5) tumors known to occur in humans and 6) the presence of
non-neoplastic lesions.
The CARC based their conclusion on the presence of uterine epithelial tumors (combined adenoma,
adenocarcinoma or malignant mixed Mullerian tumors) in female Wistar Han rats and hemangiomas
and hemangiosarcomas in male B6C3F1 mice.
In separate comments, the NTP also agreed with the CARC that TBBPA is likely to be carcinogenic to
/7umons(NIEHS, 2015).
G-8 Studies Proposed for Risk Assessment
Data on tumor incidences are available from the NTP cancer bioassay (NTP, 2014a). Based on EPA's
CARC and the NTP conclusions that TBBPA is likely to be carcinogenic to humans and because MOA
information is limited, EPA/OPPT will evaluate risks using data on uterine tumors (females) and
hemangiomas and hemangiosarcomas (males) from this study Appendix J describes the method used
to determine slope factors from a linear low-dose extrapolation. EPA/OPPT will use these slope factors
in a risk assessment of TBBPA.
EPA/OPPT will also use a study of developmental effects to assess risk for children in scenarios
discussed in Chapter 0. Although EPA/OPPT may consider other studies further, the study by Fukuda et
al. (2004) is likely to be the most appropriate study for several reasons. The study directly dosed young
animals so that matching results with exposure data is easier than studies that dosed both dams and
offspring. It is possible that TBBPA causes adverse effects primarily as a result of exposure to TBBPA by
young rodents. This is a plausible conclusion based on the breadth of studies that appear to suggest
effects are more likely when young animals are dosed directly. The European Union suggested that the
kidney effects observed in the newborn rats could be due to immature metabolic capability and/or the
immature kidneys of such young animals (EC, 2006).
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G-9 Uncertainties
G-9-1 Carcinogenesis Bioassay Methods
NTP (2014a) administered TBBPA to rodents via oral gavage throughout their lifetime. It is not known
whether similar adverse effects would result if humans are exposed to TBBPA associated with dust
particles or as other forms for less than a lifetime (or even a full lifetime).
The doses used in NTP (2014a) ranged from 250 to 1000 mg/kg-bw/day. It is possible that at lower
exposures likely to be experienced by humans, TBBPA would be excreted before it could exert adverse
effects. Yet, the available toxicokinetics data are limited in their ability to answer this question. Lower-
dose toxicokinetics studies show similar rates of excretion as higher doses via the oral route and the
longest toxicokinetics studies are 10 days (Kuester et al., 2007). Therefore, TBBPA's toxicokinetic
behavior after longer exposure durations, lower concentrations and exposure via dust (vs. gavage) is
not known.
G-9-2 Developmental Toxicity Data
The studies evaluating reproductive and developmental toxicity show a wide variety of results from no
effects up to very high doses to some subclinical effects at low doses. Also, it is not clear whether
dosing dams and offspring or just dosing offspring results in effects of TBBPA treatment. Thus, there is
uncertainty in choosing any developmental toxicity study for evaluation in a quantitative risk
assessment of TBBPA.
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Appendix H Parameters Needed for Estimating Fish Ingestion
Bioconcentration factors and fish ingestion rates are needed to estimate risks from TBBPA intake
through eating fish from one water body identified near one of the Arkansas manufacturing facilities.
Chapter 0 presents bioconcentration factors (BCFs) that can be used to estimate the concentration of
TBBPA in fish from the predicted water concentrations. For this assessment, EPA/OPPT will consider
using 720 I/kg, measured using eastern oysters (GLCC, 1989b). The values of 1200-1300 L/kg from
GLCC (1989a) are higher, but they are based on measuring total radioactivity and reflect measurement
of the parent compound and metabolites rather than the parent compound only; thus, this value is not
being considered. The lowest measured BCF is 20 L/kg (EC, 2000).
For this assessment, EPA/OPPT plans to use a fish consumption rate of 22 g/day for the recreational
fishers who may fish from this water body and the adult members of their family. This value is the
default fish consumption rate for adults recommended for the protection of human health in EPA's
Office of Water (OW) 2014 draft ambient water quality criteria (AWQC). This value represents the 90th
percentile consumption rate of freshwater and estuarine finfish and shellfish for the US adult
population 21 years and older as summarized in Table 9a of EPA (2014d). The estimate is based on data
from the 2003-2004 National Health and Nutrition Evaluation Survey (NHANES). Statewide studies of
freshwater recreational fish intake are summarized in the Exposure Factors Handbook (EPA, 2011b) for
a number of states. However, there is no study available from Arkansas. Mean intake from statewide
surveys, as reported in EPA (2011b), range from 5 to 51 g/day and 95th percentile values range from
14-61 g/day.
For children, EPA/OPPT is considering using 90th percentile consumption rates of freshwater and
estuarine finfish and shellfish, also using NHANES data, for specific age groups as detailed in Table 20a
of EPA (2014d). For children age 1 to <3 years the value is 4.7 g/day; for children 3 to <6 year the value
is 5.8 g/day; and for children age 6 to <11 year the value is 7.7 g/day.
Children < 11 is the age range most relevant for considering effects of developing cancer given that this
is a bit more than one tenth of the human life span (EPA, 2005). A shorter age range, from one < three
years, could be an appropriate age range for developmental effects based on toxicity studies
suggesting effects in newborns vs. a lack of effects in somewhat older rodents (Fukuda et al., 2004).
The body weight used in estimating intake of TBBPA via fish ingestion will correspond to the receptor
and toxicological endpoint of interest. For estimating risks of developing uterine tumors, EPA/OPPT
proposes to use the body weight of 74.8 kilograms, which is the average weight of female adults aged
30 to <40 years old as presented in the Exposure Factors Handbook (EPA, 2011b). For estimating risks
of developing hemangiomas or hemangiosarcomas, EPA proposes using a value of 87.0, the average
weight for males aged 30 to <40 (EPA, 2011b). For children aged 1 to <2 years the recommended mean
body weight is 11.4 kg; for 2 to <3 year the weight is 13.8 kg; for 3 to <6 years the weight is 18.6 kg;
and for 6 to <11 years the weight is 31.8 kg (EPA, 2011b).
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Exposure can be estimated using these parameters in the following equation:
X = (WC * BCF * FC)/(1000 * BW)
Where:
X = Intake of TBBPA via fish ingestion, u.g/kg-bw/day
WC = TBBPA water concentration, u.g/L
BCF = Bioconcentration factor, L/kg
FC = Fish consumption, g/day
1000 = Conversion factor for fish weight, g/kg
BW = Body weight of human receptors being assessed, kg
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Appendix I Draft Approach for Estimating Exposure from
Mouthing of TBBPA
TBBPA flame retardants in this cluster are either additive or reactive components used in plastic
articles and may or may not be chemically bonded to the polymers. Thus, TBBPA may migrate from
the polymer matrix to the surface of the article and have the potential for exposures through direct
contact. The presence of TBBPA in a variety of products indicates the potential for consumer exposure
via object to mouth contact and hand to mouth contact with articles. EPA/OPPT will consider the
methods and parameters discussed below to estimate mouthing exposures.
Table_Apx 1-1: Plan for Evaluating Risk to Children from Ingestion of TBBPA from Products in the
Home
Exposure Scenario
Rationale
Assessment Approach
Mouthing of products
(object-to-mouth) as well
as hand-to-mouth
transfer by young
children in the home
Exposures are expected to be
highest in children; younger
children are expected to have
longer duration of mouthing
activity when compared to older
children and adults.
Sufficient data to quantify
exposure and toxicity
Concentrations of TBBPA in products and surface
loadings will be combined with age-specific
activity patterns and exposure factors to
estimate an exposure in mg/kg-bw/day for
children. Measured or estimated migration
rates in saliva can also be considered if available.
To estimate developmental risk, exposure for 1-
year olds will be compared with a toxicity value
from a developmental study in a margin of
exposure (MOE) evaluation.
For cancer risk, exposure for children in a wider
age range (0 to 7) will be multiplied by cancer
slope factors to estimate cancer risk.
EPA/OPPT will consult with CPSC and conduct
additional literature searches to identify
whether migration data specific to TBBPA are
available.
Data from published and unpublished literature
as well as previous assessments will be
considered.
1-1
TBBPA Concentrations in Products and Product Surfaces
Recent studies have shown that TBBPA is present in many different types of consumer products and
articles as well as on the surfaces of many products. These products include electronic appliances,
electronic devices, plastic toys, plastic jewelry and tents (Di Napoli-Davis and Owens, 2013; Gallen et
al., 2014; Keller et al., 2014; Samsonek and Puype, 2013; van Bergen and Stone, 2014). Gallen et al.
(2014) evaluated the TBBPA content of many different products using X-ray fluorescence, wipe
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sampling and destructive methods to estimate concentrations. Di Napoli-Davis and Owens (2013) used
wipe sampling on the surface of electronic products, van Bergen and Stone (2014) used destructive
methods (cryogenic milling and gas chromatography/mass spectrometry) to estimate concentrations
for a wide range of products and product components. Within the Washington State database (WSDE,
2014a), TBBPA concentrations within products are reported as ranges; in the proposed TBBPA risk
assessment, EPA/OPPT could use the mid-point of these ranges with conversion factors to estimate
loading on the surface of various products. Keller et al. (2014) used wipe sampling to estimate TBBPA
loading on the surface of a tent. Samsonek and Puype (2013) measured TBBPA on the surface of black
coffee mugs. The results of these studies show that the range of TBBPA concentrations and surface
loadings in and on a variety of products span several orders of magnitude.
EPA/OPPT may use these data to estimate a range of TBBPA surface loadings (TBBPA/cm2) potentially
available for transfer to hands and into saliva.
1-2 Migration Rates into Saliva
The migration from the surface of a product into human saliva can be measured through in vivo or in
vitro testing. Results of these tests are expressed as the mass of a chemical per surface area per unit
time. For example, values could be expressed as pg, u.g or mg per either cm2 or 10 cm2 per minute.
In an in vivo study, central tendency migration rates for polybrominated diphenyl ethers (PBDEs)
ranged from 0.000003 to 0.0026 to u.g/10 cm2-minute and high-end migration rates ranged from
0.00004 to 0.03 u.g/10 cm2-minute (Chen et al., 2009). Babich (2014) summarized available in vitro
testing of phthalates; central tendency migration rates ranged from 1.1 u.g/10 cm2-mintue to 4.4 u.g/10
cm2-minute while high-end migration rates ranged from 1.9 u.g/10cm2-minute to 11.4 u.g/10 cm2-
minute (Babich, 2014).
In vitro and in vivo estimates can vary over several orders of magnitude for a given chemical and
product combination. PBDEs have very low water solubility (and lower migration rates) while
phthalates have moderate water solubility. While many other physical-chemical properties may
influence migration potential into saliva, water solubility is likely to be one of the most important
factors. Thus, because TBBPA's water solubility is between the values reported for PBDEs and
phthalates, it is possible that the migration rate of TBBPA into saliva could be between the ranges
reported in Babich (2014) and Chen et al. (2009).
1-3 Children's Activity Patterns
Based on the type of product sampled in the above studies, an object may be handled routinely by a
child, resulting in TBBPA transfer from the object to the child's hand and then subsequently to his or
her mouth. An article could also be directly placed in a child's mouth. Thus, EPA/OPPT will assign these
objects as having the potential for either "hand to mouth" transfer or "object to mouth" transfer.
Because children have higher hand-to-mouth and object-to-mouth activity than adults and adult
exposure is expected to be minimal, EPA/OPPT will evaluate exposure for only for children.
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Furthermore, mouthing behaviors are typically more prevalent among the youngest children. Thus, to
estimate developmental toxicity risks, EPA/OPPT will evaluate exposure for 1-year olds, which is the
age with the highest exposure potential as a result of mouthing behavior.
However, to estimate cancer risk, a wider age range is recommended because a longer exposure
duration is thought to be required before tumors develop. EPA/OPPT proposes evaluating exposures
for children from age 0 through 7 years old to estimate cancer risks.
EPA will use data on the frequency and duration of touching objects and placing them in mouths; this
information will be adapted from the Exposure Factors Handbook (EPA, 2011b).
1-4 Other Parameters Being Considered
Additional default parameters used in EPA's Standard Operating Procedures for Residential Pesticide
Exposure Assessment (EPA, 2012c), EPA's SHEDS model (EPA, 2008b), or contained within the Transfer
Efficiency Database (Gorman, 2012) could also be considered. Measured data is preferred over
defaults.
1-5 Method to Combine Exposure Data
The variables used to estimate exposures could be considered in a deterministic way, assigning a fixed
variable representative of high-end or central-tendency exposures (EPA, 2012c). These variables could
also be considered in a probabilistic manner by sampling from distributions (EPA, 2008b; Ozkaynak et
al., 2011).
1-6 Confidence and Uncertainty in the Available Data
A variety of the parameters to be included in any modeling are highly variable and some may be data
poor.
The potential for exposure may increase under the following circumstances: (1) if routine contact and
close proximity with the product is expected; (2) if the product contacts liquid; (3) if the product is
often warm or hot; or (4) if very young children use or are frequently near such products (Gallen et al.,
2014).
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Appendix J Dose-Response Assessment for Cancer Endpoints
The data from NTP (2014a) described in Appendix G were used to develop a dose-response
relationship using linear low-dose extrapolation.
J-l Choice of Model, Points of Departure and Oral Slope Factors
The CARC recommended a model that uses low dose linear extrapolation20 for quantification of human
risk from TBBPA (CARC, 2014). EPA/OPPT agrees that using a linear low-dose model is most
appropriate because there are limitations that preclude a clear understanding of TBBPA's MOA in
relation to the observed tumors. If TBBPA acts via a non-linear/threshold mechanism, the use of linear
low dose extrapolation will over-predict cancer risk.
EPA/OPPT modeled the dose response for uterine tumors and for hemangiomas and
hemangiosarcomas using a cancer multistage model (EPA, 2012a), with linear extrapolation at low
doses as recommended by EPA (2005) to determine the point of departure — the 90% lower bound
benchmark dose (BMDL). Using the BMDL, two slope factors were calculated for females using the
following data: 1) combined incidence of all uterine tumors of epithelial origin (adenocarcinomas,
adenomas and malignant mixed Mullerian tumors); and 2) uterine adenocarcinomas only. Likewise,
two slope factors were calculated for males: 1) the combined incidence of hemangiomas or
hemangiosarcomas; and 2) hemangiosarcomas only (Hummel, 2013b, 2014). Table_Apx J-l lists BMDs,
BMDLs and slope factors.
Based on recommendations from EPA (2005) and EPA (2012a), EPA/OPPT used the dose associated
with a 10% extra cancer risk as the point of departure to model the TBBPA tumor data.
The oral slope factor, which approximates a 95% confidence limit, is the upper bound on increased
cancer risk from a lifetime of oral exposure to a chemical and is usually expressed as a proportion of
the population affected per mg of the chemical per kg-bw/day (EPA, 2011c). The oral slope factor can
also be used to determine the extra lifetime risk that an individual may develop cancer. In the TBBPA
assessment, EPA/OPPT will use the slope factor to calculate individual risk.
The slope factor for the risk of developing tumors was calculated using the following equation:
Equation 1 SFfhuman, daily) = 0.1/(BMDL(animal, 10% extra risk) * DAF * 5/7)
Where:
SF(human, daily) = slope factor expressing the extra lifetime risk (mg TBBPA/kg-bw/day)"1
0.1 = benchmark response level (10%)
BMDL = lower 90% confidence bound on the benchmark dose
DAF = dosimetric adjustment factor = (BWanimai/BWhuman)(1/4) = 0.24
20 The CARC recommended the Ql* for modeling (CARC, 2014), but EPA/OPPT used a multistage model according to more
recent EPA recommendations (EPA, 2012a).
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5/7
= conversion from NTP (2014a) dosing of 5 days/week to 7 days/week exposure
(this factor is used only for the general population exposures)
Table_Apx J-l: Parameters Used in Dose-Response Equations
BMDL(animal, 10% extra risk)
mg/kg-bw/day
BMDL(ioHED)a
mg/kg-bw/day
Jr(human, daily)
(mg/k-bw/day)"1
Tumors in Females
Uterine tumors
(combined)
Uterine
adenocarcinomas
177
191
42.6
45.9
0.00329
0.00305
Tumors in Males
Hemangiosarcomas
and hemangiomas
Hemangiosarcomas
216
200
51.8
48.0
0.00270
0.00292
Sources: Hummel (2013b); Hummel (2014)
BMDL(iOHED) = BMDLfanimal, 10% extra risk) * DAF
Further work will be needed to determine factors to account for the amount of inhaled particulates
that might be swallowed/absorbed. The European Union suggested that 70% of inhaled particulates
might be swallowed (EC, 2006). EPA/OPPT has not yet determined a value for the proposed
assessment.
J-2 Calculation of Target Risk Levels
The slope factor can be used to calculate a TBBPA dose associated with three target risk levels:
SF(human,daily)
Equation 2
(1 mg TBBPA per kg per bw per day)
RL
x
Where:
SF(human, daily) = extra lifetime risk per 1 mg TBBPA/kg-bw/day from Eq. 1
RL = target risk level (1/10,000; 1/100,000; or 1/1,000,000)
x = the daily dose associated with the chosen target risk level (mg/kg-bw/day)
The relationship can be rearranged to solve for x for any target risk level:
RL*1 mg TBBPA per kg per bw
X =
SF(human,daily)
Resulting doses associated with the three risk levels are listed in Table_Apx J-2.
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Table_Apx J-2: Doses (in mg/kg-bw/day) Associated with Three Target Risk Levels
Uterine tumors
(combined)
Uterine
adenocarcinomas
Hemangiosarcomas
and hemangiomas
Hemangiosarcomas
Target Risk Level
1 x ID'6
0.000304
0.000328
0.000370
0.000343
1 x ID'5
0.00304
0.00328
0.00370
0.00343
1 x ID'4
0.0304
0.0328
0.0370
0.0343
J-3 Key Sources of Uncertainty
J-3-1 Tumors Modeled
NTP concluded that for hepatoblastomas in male mice, there is some evidence of carcinogenic activity
that can be attributed to TBBPA. Thus, EPA/OPPT investigated possible dose-response relationships
using the cancer multistage model and other models. The effects at 250 and 500 mg/kg-bw/day were
statistically significantly different from controls using pairwise comparisons (p < 0.05). Furthermore,
the data showed a statistically significant dose-related trend using the one-sided Cochran-Armitage
trend test (p < 0.05). However, the data did not provide a good fit using the cancer multistage model
(Hummel, 2013b), and none of the other models available in the benchmark dose response modeling
software resulted in goodness of fit p-values at acceptable levels of > 0.1 when considering the full
shape of the dose-response curve (Hummel, 2013a). Thus, this tumor type cannot be considered in a
quantitative risk assessment of TBBPA.
J-3-2 Use of Linear Low-Dose Extrapolation
The CARC noted that according to the IPCS MOA framework (IPCS, 2007), data are not adequate to
draw conclusions about the mode(s) of action for the tumor incidence associated with TBBPA.
Therefore, as recommended by EPA (2005), linear low dose extrapolation was used as the default
option for modeling tumor data.
There are several reasons that a non-linear mode of action could explain the relationship between
TBBPA and tumor incidence. EPA/OPPT and the CARC determined that a direct genotoxic MOA for
tumorigenesis is unlikely. In addition, uterine tumors might be a result of TBBPA's competitive binding
of enzymes involved in the conjugation of endogenous estrogens or TBBPA's enzyme inhibition; any
resulting higher estrogen levels might lead to tumors only after a threshold dose of TBBPA is achieved.
Furthermore, if damage to DNA occurs from the generation of reactive oxygen species (e.g., as a result
of metabolism of TBBPA), such damage may also lead to tumor formation via a non-linear MOA.
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J-3-3 Choice of Dose-Response Model
Cancer multistage models were chosen for modeling based on adequate fits for two tumor types and
biological considerations even though other models also resulted in adequate fits of the data.
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Appendix K EU and Canada Risk Assessments: Specific
Evaluations Consulted for Current Assessment
Some of the more detailed evaluations of the EU and Canadian risk assessments that were considered
during problem formulation are described below.
K-l EU Calculations of TBBPA Uptake from Soil Near
Manufacturing and Processing Sites
The European Union (EU) human health risk assessment of TBBPA (EC, 2006) estimated risks used a log
Kow of 5.9 and a Koc of 49,726 L/kg to calculate uptake of TBBPA from soil to plants and livestock for
several manufacturing (and processing) scenarios. Resulting estimates of dietary intake ranged from
1.7 x 10"5 to 2.92 mg/kg-bw/day, with the highest value estimated at sites where TBBPA is used as an
intermediate in the production of TBBPA derivatives. The next highest estimate was 2.33 mg/kg-
bw/day for individuals living near manufacturing facilities.21 EPA/OPPT does not plan to use log Kow or
Koc values based on the potential for predict uptake that is higher than actual uptake to
plants/livestock.
K-2 EU Environmental Assessment at Processing Sites
The EU risk assessment for the environment evaluated processing sites and a discussion of their
analysis and applicability to the current TBBPA assessment is described below.
K-2-1 Acrylonitrile Butadiene Styrene (ABS) Compounding
Sites
Although the EU found risks at processing sites, the generic scenarios used in the EU risk assessment
for ABS compounding facilities would not readily apply to United States facilities that process TBBPA
because reported release information from TRI differs from the release assumptions used by the EU.
Using two approaches, the EU concluded that environmental risks are possible at ABS (plastic)
compounding sites that use TBBPA as an additive flame retardant (EC, 2008):
• a generic scenario evaluation based on flame retardant use in polymers with some TBBPA-
specific data using certain worst-case assumptions and
• a site-specific evaluation using monitored emissions data from a single site.
For the generic approach, the EU used default release estimates and several worst-case assumptions.
The largest releases were TBBPA dust losses from raw materials handling and other particulate
releases, which eventually went to wastewater.
21These two highest values reported in the EU risk assessment were used only as examples because the EU doesn't produce
TBBPA or its derivatives (EC, 2006).
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In addition to these more generic scenarios, the EU monitored TBBPA emissions from one ABS
compounding site and identified risks for aquatic, sediment- and soil-dwelling organisms. The EU
specifically stated that air emissions led to risks for soil-dwelling organisms. The site-specific
monitoring information is confidential, however. Therefore, the amount of TBBPA emitted and the
types of emissions (i.e. whether there were also releases to land and water in addition to air releases)
are not known (EC, 2008).
For all years of TBBPA reporting to TRI, the largest amount transferred to WWTPs occurred in 2003,
when 152 pounds (all facilities) were released to wastewater. In nine of the reporting years (2000-
2002; 2007-2012), 10 or fewer pounds of TBBPA were sent to wastewater treatment plants (WWTPs).
Facilities reported minimal direct releases to surface water. The highest value was 71 pounds in 2003
(EPA, 2012e). Thus, assumptions related to releases differ significantly between the European Union
and the United States.
In the United States during 2012, six facilities were identified in the category of plastics and rubber
facilities (NAICS code 326), the category that would include ABS compounding sites. One of these
facilities reported 33 pounds as fugitive air emissions. Four of the facilities disposed a total of 154
pounds of TBBPA to "other landfills." No other releases (including stack air releases) were reported for
these facilities. Similar emissions were reported for facilities within this industry category in previous
years. Between 2000 and 2012, the highest fugitive air emissions for a single facility was 37 pounds, in
2010 (EPA, 2012e). Thus, for similar types of processing plants as those evaluated by the EU, only
limited emissions have been reported to TRI for US facilities.
Overall, both the types and amounts of emissions reported for these facilities differ between the EU
and the United States. Unlike the releases reported in the EU analysis for the ABS compounding sites,
releases reported by US plastics and rubber facilities to TRI are not expected to result in environmental
risks.
K-2-2 ABS Conversion Sites and Epoxy Resin Manufacturing
Facilities
The EU risk assessment also found risks for soil-dwelling organisms surrounding facilities that
manufacture epoxy and/or polycarbonate resins and at ABS conversion sites. However, these risks
were found for situations where sludge is generated/applied to agricultural land. The EU notes that
when these activities were taken into account, risks were not identified for these generic scenarios.
Furthermore, the EU did not identify environmental risks when using actual site-specific data at eight
epoxy resin manufacturers and at two sites using TBBPA in reactive flame retardant applications (EC,
2008).
US facilities have reported minimal TBBPA releases to surface water and WWTPs as stated above. Also,
no releases were reported to land treatment (EPA, 2012e).
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K-2-3 EPA's Conclusion for Processing Sites
When considering ABS compounding sites (reactive uses), TRI reporting shows that plastics and rubber
facilities reported minimal air emissions and no other releases of note.
Other processing sites in the United States have reported higher stack air releases. However, given that
these are a small proportion of the air emissions from the top manufacturing site, EPA/OPPT did not
further evaluate these releases.
For manufacturing epoxy and polycarbonate resins and ABS conversion sites, measured data for the
ten surveyed sites in the EU risk assessment showed low risks. Although some risks were identified in
generic scenarios in the EU risk assessment, TRI data indicate minimal or no releases from such
processing sites in the United States.
K-2-4 Canada's Fugacity-Based Model of Sludge Applied to Land
The Canadian government (EC/HC, 2013) ran a fugacity-based model to estimate risks from TBBPA in
sludge applied to land. Using different assumptions from those used in EPA/OPPT's preliminary
calculation (but similarly using a high-end TBBPA sludge concentration), risks to soil organisms were
determined to be low (EC/HC, 2013). Although EPA/OPPT doesn't use these fugacity-based models
when estimating exposure, these results support EPA/OPPT's preliminary calculation for this pathway.
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